Biodegradable plastics in the marine environment: a potential source of risk?
Abstract
The marine environment is facing the threat of increasing plastic pollution, especially from disposable plastics. Presently, governments worldwide are promoting policies to restrict or prohibit conventional plastics. As one hopeful alternative to conventional disposable/non-durable plastics, biodegradable plastics have attracted much attention and controversy in terms of their definition, environmental impact, and environmental significance, as they may be widely used. Therefore, it is necessary to clarify the facts about biodegradable plastics, understand the current knowledge gaps, and identify promising fields of relevant research. This review briefly introduces some common biodegradable plastics, their mechanisms of biodegradation, indicators for the biodegradation process, and factors concerning biodegradability and summarizes studies on the biodegradation of biodegradable plastics in the marine environment. The lifespan of biodegradable plastics varies greatly due to their compositions/properties as well as significant differences in the marine environment. Then, the potential risks of biodegradable plastics, including the release of pollutants (micro/nanoplastics, degradation products, and additives), adsorption-desorption of pollutants (pesticide, etc.), and their impact on biomes and biogeochemical cycles are discussed, fully revealing their possible impacts on the marine environment. It is believed that, in addition to the waste of resources, a high abundance of microplastics, toxic leachates, and complex effects on habitats and the environment may also cause problems for the marine environment as a result of the widespread and inappropriate use of biodegradable plastics. Based on the discussion, some constructive suggestions on how to use biodegradable plastics reasonably and prudently in the future are put forward.
Keywords
INTRODUCTION
Since the last century, human life has been deeply affected by artificial polymers, such as plastic and rubber, as well as the pollutants from their products[1] (marine plastic litter, microplastics, plasticizers, etc.). With the widespread concerns and appeals of various sectors of society[1], new solutions are emerging to reduce artificial polymer pollution[2]. As one of the most promising options, biodegradable plastics (BPs) are widely considered an alternative to conventional plastics in Europe, the USA, and China[3]. BPs are artificially synthesized or chemically modified polymers that can be completely mineralized[4], i.e., eventually converted into carbon dioxide and water, by biological activity[5] or, in a narrow sense, naturally occurring microorganisms[6]. Similar and confusing nouns include “bioplastics” and “bio-based plastics”. Different from bio-based plastics, which are wholly or partially made from biological materials and not necessarily compostable or biodegradable, BPs could be made from fossil fuel-based materials. As a less formal term, “bioplastics” is sometimes used as a collective term to cover both bio-based plastics and BPs[7] or specifically to refer to bio-based plastics[8] [Table 1].
Common BPs: their sources, properties, and usage
| Composition | Abbreviation | Source* | Material density | Usage§ |
| Poly(lactic acid) | PLA | Bio-based | 1.25 | Packaging, paper coatings, mulch films, compost bags, etc. |
| Polyhydroxyalkanoate | PHA | Bio-based | 1.20-1.25 | Disposable drinking cups, cutlery, trays, food plates and food containers; soil retention sheathing and other agricultural films; garbage bags, shopping bags, etc. |
| Polyhydroxybutyrate | PHB | Bio-based | 1.18-1.26 | Bottles, bags, packaging film and disposable nappies, etc. |
| Poly(hydroxybutyrate-co-valerate) | PHBV | Bio-based | 1.25 | Films and paper coatings, biomedical applications, etc. |
| Starch blends | - | Bio-based | 1.25-1.35 | Food packaging, bottles, cutlery, straws, disposable bags, etc. |
| Polyglycolic acid | PGA | Bio-based | 1.53 | Packaging films for oxygen-sensitive products, shale gas mining and other industrial processes, and synthetic fast absorbable sutures for surgery |
| Cellulose acetate | CA | Bio-based | 1.20-1.30 | Textiles, optical film for LCD technology, antifog goggles, filters, etc. |
| Poly(ε-caprolactone) | PCL | Fossil-based | 1.12 | Mulch and other agricultural films, etc. |
| Poly(butylene Succinate) | PBS | Bio-/fossil -based | 1.23-1.26 | Food packaging, mulch film, plant pots, hygiene products, fishing nets, fishing lines, etc. |
| Poly(butylene succinate-co-butylene adipate) | PBSA | Bio-/fossil-based | 1.23 | Ditto |
| Poly(butylene adipate) | PBA | Fossil-based | 1.02-1.12 | Injection molding for automotive, mechanical and electronic industries, etc. |
| Poly(butylene adipate-co-terephthalate) | PBAT | Fossil-based | 1.25 | Food packaging, agricultural film, etc. |
BPs are primarily used as disposable or non-durable plastic products, such as packaging, mulching film, and tableware[16]. These items accounted for 49% of global plastic production in 2019[17]. As early as 2018, before the COVID-19 pandemic, 5% of annual municipal solid waste and 40% of plastic waste in the United States were containers and packaging plastics, as estimated by USEPA (by weight)[18,19]. It is notable that only 14.5% of this waste was recycled in the United States[18]. For these products, the short service time greatly increases the frequency of their discovery in the environment. For example, disposable or non-durable plastic products are common in marine litter globally[20]. Moreover, the pandemic has led to increased demand for single-use plastics, including plastic bags, electronic packaging, food delivery and takeout packaging, and personal protective equipment[21]. The International Solid Waste Association shows that consumption of single-use plastics has increased by 250%-300% since the pandemic began[17], accompanied by secondary pollutants, such as ubiquitous microplastics in the marine environment, from abandoned personal protective equipment[22]. Although the pandemic will end one day, its impact on consumption habits (such as online shopping and express delivery, in which plastic packaging is widely used) might continue for a long time; meanwhile, it reminds us that disposable plastic products cannot be completely replaced[21]. Especially in recent years, the successive restrictions or bans on the use of single-use plastics put forward by a variety of countries have aroused a huge demand for BPs[2,3]. Considering that BPs accounted for only 0.3% and 0.6% of plastic production in 2019[23] and 2020[24], respectively, the consumption of BPs should increase greatly in the near future, as predicted. The consumption of “biodegradable and compostable plastic products” in Europe was estimated to be beyond 300,000 tons in 2020, three times the consumption in 2015[25].
This does not mean that there is no controversy in novel polymer materials known as “biodegradable plastics”. One of the most common questions is whether BPs can degrade efficiently in the real and complex natural environment[9], including the aquatic environment. Firstly, the definition of “biodegradable” is vague[23]. Actually, it should be expressed as “biodegradable under certain conditions”[3]. According to the retrieval of American Standard for Testing and Materials (ASTM) standards of test methods [Table 2], BPs could be divided into several categories, including “industrial compostable”, “marine biodegradable”, and “soil biodegradable”, which are also certified in the European certification scheme[3]. There is also another relatively sound standard system for BPs, the International Organization for Standardization (ISO) standard system [Table 2]. Most certified “biodegradable plastics” could only be efficiently biodegraded in aerobic or industrial anaerobic conditions. The relatively high temperature and ideal moisture content in industrial composting could accelerate the biodegradation process; it is also suggested that the aquatic environment has fewer microbes per unit than compost and soil environments[27]. Consequently, the average biodegradation level of BPs in industrial composting is 72.3% over 75 days, based on the data obtained from comprehensive data analysis, while that in the marine environment is 47.1% over 155 days[27]. In addition, there are unignorable differences between laboratory test conditions and the natural environment[28]. Therefore, the degradability of BPs might be overestimated. For example, the material densities of most BPs [Table 1] are higher than that of surface water (including seawater); thus, immediately after entering the aquatic environment, they could sink to the bottom[29], where the environment may be dark with anoxic or anaerobic conditions and maintain a relatively low temperature. There must be deviations and uncertainties between the supposed biodegradation rate and the actual rate in the environment.
Active American Standard for Testing and Materials (ASTM) and International Organization for Standardization (ISO) standards of test methods concerning “biodegradable plastics”[26]
| Classification | Environment | Degradation pattern | Standard number | Standard name |
| ASTM standards | ||||
| Solid waste disposal site | Landfill | Anaerobic | D5526-18 | Standard test method for determining anaerobic biodegradation of plastic materials under accelerated landfill conditions |
| Aerobic and/or anaerobic | D7475-20 | Standard test method for determining the aerobic degradation and anaerobic biodegradation of plastic materials under accelerated bioreactor landfill conditions | ||
| High-solids anaerobic digester or biologically active landfill | Anaerobic | D5511-18 | Standard test method for determining anaerobic biodegradation of plastic materials under high-solids anaerobic-digestion conditions | |
| Composting | Aerobic | D5338-15(2021) | Standard test method for determining aerobic biodegradation of plastic materials under controlled composting conditions, incorporating thermophilic temperatures | |
| Aerobic | D6400-21 | Standard specification for labeling of plastics designed to be aerobically composted in municipal or industrial facilities | ||
| Natural environment | Soil | Aerobic | D5988-18 | Standard test method for determining aerobic biodegradation of plastic materials in soil |
| Marine environment | Aerobic | D6691-17 | Standard test method for determining aerobic biodegradation of plastic materials in the marine environment by a defined microbial consortium or natural sea water inoculum | |
| Aerobic | D7991-15 | Standard test method for determining aerobic biodegradation of plastics buried in sandy marine sediment under controlled laboratory conditions | ||
| Either | Soil, landfill, compost, land cover and agricultural use | Abiotic+ Aerobic/anaerobic | D6954-18 | Standard guide for exposing and testing plastics that degrade in the environment by a combination of oxidation and biodegradation |
| ISO standards | ||||
| Solid waste disposal site | High-solids anaerobic digester | Anaerobic | 15985:2014 | Plastics - Determination of the ultimate anaerobic biodegradation under high-solids anaerobic-digestion conditions - Method by analysis of released biogas |
| Composting | Aerobic | 14855-1:2012 | Determination of the ultimate aerobic biodegradability of plastic materials under controlled composting conditions - Method by analysis of evolved carbon dioxide - Part 1: General method | |
| 14855-2:2018 | Determination of the ultimate aerobic biodegradability of plastic materials under controlled composting conditions - Method by analysis of evolved carbon dioxide - Part 2: Gravimetric measurement of carbon dioxide evolved in a laboratory-scale test | |||
| 16929:2021 | Plastics - Determination of the degree of disintegration of plastic materials under defined composting conditions in a pilot-scale test | |||
| 20200:2015 | Plastics - Determination of the degree of disintegration of plastic materials under simulated composting conditions in a laboratory-scale test | |||
| Natural environment | Soil | Aerobic | 17556:2019 | Plastics - Determination of the ultimate aerobic biodegradability of plastic materials in soil by measuring the oxygen demand in a respirometer or the amount of carbon dioxide evolved |
| Aqueous medium | Aerobic | 14851:2019 | Determination of the ultimate aerobic biodegradability of plastic materials in an aqueous medium - Method by measuring the oxygen demand in a closed respirometer | |
| Aerobic | 14852:2021 | Determination of the ultimate aerobic biodegradability of plastic materials in an aqueous medium - Method by analysis of evolved carbon dioxide | ||
| Anaerobic | 14853:2016 | Plastics - Determination of the ultimate anaerobic biodegradation of plastic materials in an aqueous system - Method by measurement of biogas production | ||
| Marine environment | Aerobic | 22766:2020 | Plastics - Determination of the degree of disintegration of plastic materials in marine habitats under real field conditions | |
| Aerobic | 22403:2020 | Plastics - Assessment of the intrinsic biodegradability of materials exposed to marine inocula under mesophilic aerobic laboratory conditions - Test methods and requirements | ||
The longer (than expected) environmental degradation term might greatly reduce the difference between BPs and conventional plastics, for example, in some risks[30], while the shorter lifespan might bring different environmental impacts than conventional plastics. Hence, the question is: In what way will BPs, which are expected to be used widely, affect the marine environment? Could they become potential sources of pollutants or carriers of harmful microorganisms in the marine environment? During degradation, will they cause secondary ecological risks, via the release of toxics to or other unpredictable impacts on the environment or ecological equilibrium, especially in the context of large-scale disposable use? This review provides some information about BPs and collects studies and views on the degradation of BPs in the marine environment and their ecological effects to deepen the understanding of the environmental impact of BPs. Significantly, some results in other environments (e.g., in soil) are also included to make inferences, since some evidence for BPs in the marine environment is still not sufficient.
MAIN TEXT
Biodegradation ability in the marine environment and related factors
Biodegradation process and description indicators
In general, the biodegradation process of BPs can be divided into three steps [Figure 1]: fragmentation, hydrolysis, and biodegradation of products (such as oligomers and monomers). (I) In the first step, processes dominated by abiotic factors or biotic factors (also named biodeterioration[31]) induce fragmentations. Abiotic factors, including mechanical stress (e.g., caused by waving scouring, sand abrasion, or animals[32]), light[33], thermal, and chemical factors[34], play their roles in the fragmentation of BPs. (II) Hydrolysis, as a depolymerization reaction, is the rate-limiting step in the biodegradation of BPs[35]. The hydrolysis rate of the whole BP could be either approximately uniform (bulk erosion mechanism) or determined by the process on the surface (surface erosion mechanism)[31]. Most BPs are aliphatic polyesters or contain glycosidic bonds [Table 1]: PLA, PBAT, and starch-based materials (e.g., Mater-bi® and Bioflex®) are among the most popular BPs[36,37] [Figure 2[7,38-41]]. They are “fast” hydrolyzed by extracellular enzymes[42], and some of them could be hydrolyzed in an alkali environment without any enzyme[43]. However, it is difficult for most hydrolases to permeate the structure of BPs[44], so bulk hydrolysis rates of BPs are determined by the bulk diffusion rate and migration distance of water molecules[45]. (III) The last step is the degradation of hydrolysis products. The products could be taken in and then mineralized by microbial cells. Therefore, this step can be divided into bioassimilation and mineralization[31]. The entire process of this step does not need to take place inside microorganisms. For example, anaerobic microorganisms can degrade BPs to CH4 and then finally mineralize them in vitro. It is worth mentioning that some believe that deterioration processes where plastics begin to lose their physical/structural properties should be the first step of biodegradation[9]. However, deterioration does not significantly change the environmental behavior of BPs. Considering this process as a part of biodegradation lacks environmental significance.
Figure 1. Diagrammatic sketch of biodegradation process of biodegradable plastics.
These three steps do not necessarily need to be performed sequentially; for example, the first step (fragmentation) may be the result of the second step (hydrolysis)[46]. Thus, some studies suggest that the biodegradation process can be summarized into two steps, namely, depolymerization and utilization of byproducts[47]; fragmentation is regarded as a pre- or simultaneous biotic/abiotic process[6]. In any case, the results of the previous step can greatly speed up the next step. For example, the fragmentation of BPs could greatly accelerate the hydrolysis reaction, especially when surface erosion plays a major role.
Several indicators, such as weight loss, molecular weight reduction, decrease in biochemical oxygen demand (BOD), and CO2/CH4 produced, have been used to describe the biodegradation of BPs [Table 3]. It is difficult to accurately describe the entire biodegradation process with a single indicator. Among them, weight loss is the most widely used [Table 3], even though it cannot track the biodegradation of secondary microplastics formed in the BP fragmentation, which rapidly disappear in the natural environment[48]. Sometimes, the photogrammetric method is used to determine the approximate weight loss[59]. Weight loss could occur throughout the biodegradation process[35] and is affected by several interference factors, such as the release of soluble additives and the loss of fragments. Molecular weight reduction, mainly caused by the chain shortening, is a representative indicator for the hydrolysis of BPs, since only in hydrolysis can there be a measurable and significant molecular weight reduction. Due to this, a greater weight loss ratio than molecular weight reduction indicates that surface erosion predominates, such as that for PCL[50,52] and PHA[47]. A contrary situation indicates that bulk erosion is more important than surface erosion at this stage, as happens to PBS, PBAT[52], and PLA[50].
Test results for biodegradation ability of BPs in the marine environment
| BPs | Shape and size | Condition | Period | Indicators for degradation | Refs. | |||
| Weight loss | Decrease in BOD | Decrease in MW* | CO2 produced | |||||
| PLA | Spline, 80 × 4 × 2 mm | Coastal seawater | 52 weeks | 2% | - | - | - | [35] |
| Film, 12 × 12 × 0.32 mm | Artificial seawater, 25 °C | 1 year | 0 | - | - | - | [48] | |
| Film, 150 × 150 × 0.05 mm | Artificial seawater | 6 months | < 1% | - | - | - | [49] | |
| Film, 30 × 3 × 0.05 mm | Static seawater, 25 °C | 10 weeks | 0 | - | 4.5% | - | [50] | |
| Film, 30 × 3 × 0.05 mm | Coastal seawater, 19-26°C | 5 weeks | 25% | - | - | - | [35] | |
| Sheet, 20 × 20 × 0.2 cm | ASTM D6691 | 56 days | - | - | - | 7%† | [51] | |
| Spline, 25 × 6 × 2 mm | Natural seawater; static seawater (lab) | 1 year | < 3% | - | 13% (natural); 25% (lab) | - | [52] | |
| Film, φ50 × 0.02 mm | On natural sediment | 82 days | 0.2% | - | - | - | [53] | |
| Film | Buried in sand | 71 days | 0.15% | - | - | - | [54] | |
| PLA mixture | Film | Seawater; sediment | 365 days | 5%-29% (seawater); 65%-99% (sediment) | - | - | - | [55] |
| PHB | Film, 12 × 12 × 0.32 mm | Artificial seawater, 25 °C | 1 year | 6% | - | - | - | [48] |
| Film, 30 × 20 × 0.1 mm | Coastal seawater | 6 weeks | 40%-100% | - | - | - | [35] | |
| Film, 0.005 mm thickness; pellet, φ10 × 5 mm | Tropic seawater | 160 days | 42% (film); 38% (pellet) | - | 26% (film); 20% (pellet) | - | [47] | |
| Film, 30 × 3 × 0.05 mm | Static seawater, 25°C | 10 weeks | 8% | - | 2.5% | - | [50] | |
| Film, 30 × 3 × 0.05 mm | Coastal seawater, 19-26°C | 5 weeks | 65% | - | - | - | [35] | |
| Film | Static seawater, 25°C | 2 weeks | 80% | - | - | [56] | ||
| Film, 0.1 mm thickness | Seawater in lab (bay and ocean), 25 °C | 28 days | 41% (bay); 23% (ocean) | 27% (bay); 14% (ocean) | - | - | [57] | |
| Film, 0.16 mm thickness | Lab/aquarium incubation; ASTM 6691 | 49 d (lab or aquarium) | Lab: 99%; 86%§. Aquarium: 30%; 72%§ | - | - | 80%-90% (100 d) | [58] | |
| Sheet, 20 × 20 × 0.2 cm | ASTM D6691 | 43 days | - | - | - | 89%† | [51] | |
| Film, 60 × 80 × 0.085 mm | Marine sediment | 356 days | 81%-97% | - | - | - | [59] | |
| Film | Buried in sand | 71 days | 15% | - | - | - | [54] | |
| PHBV | Film, 0.005 mm thickness; pellet, 10 mm diameter, 5 mm height | Tropic seawater | 160 days | 46% (film); 13% (pellet) | - | 16% (film); 57% (pellet) | - | [47] |
| Fiber, 0.26 mm diameter | Seawater | 8 weeks | 65% | - | - | - | [35] | |
| Film, 0.1 mm thickness, P(3HB-co-14%3HV) | Seawater in lab (bay and ocean), 25 °C | 28 days | 100% | 84% (bay); 78% (ocean) | - | - | [57] | |
| Film, 0.29 mm thickness | Lab/aquarium (aqu) incubation; ASTM 6691 | 49 d (lab/ aqu) | Lab: 99%; 87%§. Aqu: 32%; 48%§ | - | - | 85%-95% (100 d) | [58] | |
| PHB-co-PHV¶ | Film, 150 × 150 × 0.05 mm | Artificial seawater | 6 months | < 1% | - | - | - | [49] |
| PHO¶ | Sheet, 20 × 20 × 0.2 cm | ASTM D6691 | 56 days | - | - | - | 38%† | [51] |
| Mater-Bi¶ | Film, 200 × 20 × 0.025 / 60 × 80 × 0.012 mm | Marine sediment | 298 d /356 days | 5%-99% /47%-98% | - | - | - | [59] |
| Carrier bags | Littoral marsh | 90 days | 1.5%-1.6% | - | - | - | [60] | |
| Carrier bags | Seawater | 90 days | 1.7%-4.5% | - | - | - | [60] | |
| TPS¶ | Sheet, 20 × 20 × 0.2 cm | ASTM D6691 | 28 days | - | - | - | 92%† | [51] |
| PCL | Film, 12 × 12 × 0.32 mm | Artificial seawater, 25 °C | 1 year | 1% | - | - | - | [48] |
| Film, 30 × 20 × 0.1 mm | Seawater | 6 weeks | 98% | - | - | - | [35] | |
| Film, 30 × 3 × 0.05 mm | Static seawater, 25 °C | 10 weeks | 24% | - | 8.5% | - | [50] | |
| Film, 30 × 3 × 0.05 mm | Coastal seawater, 19-26 °C | 5 weeks | 34% | - | - | - | [35] | |
| Film, 0.1 mm thickness | Seawater in lab (bay and ocean), 25 °C | 28 days | 100% (bay); 67% (ocean) | - | - | - | [57] | |
| Spline, 80 × 4 × 2 mm | Coastal seawater | 52 weeks | 30% | - | - | - | [35] | |
| Sheet, 20 × 20 × 0.2 cm | ASTM D6691 | 56 days | - | - | - | 80%† | [51] | |
| Spline, 25 × 6 × 2 mm | Natural seawater; static seawater (lab) | 1 year | 32% (natural); 12% (lab) | - | 1% (natural); 7% (lab) | - | [52] | |
| Film | Buried in sand | 71 days | 1.8% | - | - | - | [54] | |
| PBS | Spline, 80 × 4 × 2 mm | Coastal seawater | 52 weeks | 2% | - | - | - | [35] |
| Film, 30 × 20 × 0.1 mm | Seawater | 6 weeks | 2% | - | - | - | [35] | |
| Film, 0.1 mm thickness | Seawater in lab (bay and ocean), 25 °C | 28 days | 2% | 1% (bay); 2% (ocean) | - | - | [57] | |
| Sheet, 20 × 20 × 0.2 cm | ASTM D6691 | 56 days | - | - | - | 21%† | [51] | |
| Spline, 25 × 6 × 2 mm | Natural seawater; static seawater (lab) | 1 year | < 3% | - | 28% (natural); 46% (lab) | - | [52] | |
| PBSA | Film | Buried in sand | 71 days | 6% | - | - | - | [54] |
| PBA | Film, 0.1 mm thickness | Seawater in lab (bay and ocean), 25 °C | 28 days | 34% (bay); 11% (ocean) | 20% (bay); 10% (ocean) | - | - | [57] |
| PBAT | Spline, 80 × 4 × 2 mm | Coastal seawater | 52 weeks | 56% | - | - | - | [35] |
| Film, 30 × 20 × 0.1 mm | Seawater | 6 weeks | 7% | - | - | - | [35] | |
| Spline, 25 × 6 × 2 mm | Natural seawater; static seawater (lab) | 1 year | < 3% | - | 56% (natural); 61% (lab) | - | [52] | |
| Film, φ50 × 0.02 mm | Seawater; on natural sediment | 82 days | 1.2%-1.5% | - | - | - | [53] | |
Indicators associated with specific compositions or products, including mechanical properties and highly ordered structures[50], are rarely used in the assessment of the natural biodegradation of BPs. On the one hand, they are not suitable for comparison between different compositions. On the other hand, some of them are too sensitive to the environment[35] and vary significantly in the deterioration process before fragmentation[9] and during the whole biodegradation process[35].
Factors affecting biodegradability
Properties of BPs
Similar to other organic substances, the biodegradability of BPs first depends on their chemical compositions. Firstly, the properties of BPs affect the adhesion of microorganisms. Properties of polymers, including surface charge, hydrophilicity/hydrophobicity, roughness, and bioavailability (biodegradability), were supposed to affect surface biofilm formation[24]. The positive surface charges on BPs (e.g., PLA) are prone to adhere to microorganisms, since zero charge points of most bacterial cells are between pH of 2 and 4[24,61]. The hydrophobicity of plastic surfaces might provide plastics with different microbial compositions from those on glass[24], ceramics[62], or the natural environment[24]. Generally, there is stronger adhesion between bacteria and a hydrophobic surface (than a hydrophilic one)[63]. The specific surface area of plastics could be affected by their size, shape, and roughness. It takes longer for marine microorganisms to come into contact with and adapt to plastics with small surface areas[47]. The specific surface area of BPs increases in the biodegradation process, with the decrease of size and the change of roughness[24].
Secondly, one of the impacts of BPs’ properties on their biodegradability, which may be the most important, is that they determine the hydrolysis rate. Some properties that proved important include the following.
Finally, some materials (PBAT, PHBV, etc.) could act as the sole carbon source for bacteria, while others (PLA and PBS) could not[67], which might be caused by the difference in the uptake ability of microorganisms to hydrolysates.
Additives
Additives are chemical substances providing plastic desired qualities[68]. Plastic additives are complex and diverse; the specific ingredients of plastic additives are often undisclosed as trade secrets. The main essential additives of BPs include crystalline nucleating agents, chain extenders, anti-hydrolysis stabilizers, melt enhancers, plasticizers, etc. Inorganic additives such as inorganic nucleating agents might cause the degradation of plastic products due to their poor compatibility. As the name implies, chain extenders could extend the chain, increase the molecular weight, and increase crosslinking during production, which prolongs the biodegradation time[44]. Anti-fouling agents[69] and chain extenders[70] could inhibit extracellular enzyme activity or prevent the microorganism from producing enzymes. Anti-hydrolysis stabilizers for condensed BPs can significantly improve their aging stability. Therefore, the leaching of plasticizers is supposed to make the plastics brittle and stimulate fragmentation[71]. Some other additives, such as colorants[72] and photostabilizers[73], could influence the final thermal and UV stability of plastics, which might affect the fragmentation and hydrolysis of BPs.
Natural environmental condition
Natural environmental conditions, including habitats and climate zones, could lead to some differences in the degradation of BPs[74]. The main environmental factors in the marine environment consist of light, temperature, salinity, pH, nutrients, electron acceptors, grain size of sediments, and microbial population. These factors can not only affect microbial composition, abundance, and activity[30] but also cause some abiotic effects on degradation.
UV light accelerates the fragmentation of BPs[71] and slowly produces some low molecule weight oxidation products for bioassimilation[30], but the effect of UV light is rapidly attenuated with water depth[74]. Light might also decelerate biodegradation, e.g., crosslinks formed during photoirradiation-induced reactions could slow down biodegradation[30]. In addition, light can moderate the proportion of photosynthetic autotrophs. Although no significant mineralization effect of plastics is expected due to their metabolic pathways[75], algae are important participants of the “plasticsphere” in the marine environment and affect BPs’ degradation indirectly.
In the marine environment, temperature, pH, and salinity change in a certain range, and ultra-high temperatures could accelerate the oxidative degradation[76] and abiotic hydrolysis[30] of plastics, while ultra-low temperatures might cause frozen-accelerated degradation[77]. However, for biodegradation, a temperature within a suitable range is needed for both the microbial growth on the surface of BPs and the reaction rate of enzymes[78]. Different microorganisms (psychrophilic, psychrotrophic, and mesophilic) thrive and are active at various water temperatures[30]. Similarly, the effects of pH and salinity on biodegradation indirectly affect the microbial communities, and different microorganisms/enzymes prefer different pH and salinity[30]. Specifically, relatively warm and alkali environments are suitable for the enzyme hydrolysis of BPs[44]. Additionally, high salinity makes abiotic hydrolysis in seawater slower than in fresh water, possibly due to the slower diffusion of seawater in polyester[35].
The availability of nutrients and electron acceptors is a key factor in the biodegradation of BPs. Essential nutrients and electron acceptors are necessary for the survival and metabolism of microorganisms. Oxygen is an important electron acceptor in the biodegradation (oxidation reaction) of BPs, but it is certainly not sufficient in the marine environment. Given the specific density of BPs, they are more likely to appear under anaerobic or anoxic conditions when offshore. Sunken ships that have not been decayed for hundreds of years can laterally illustrate the impact of the hypoxic environment on biodegradation[30]. However, even in this case, biodegradation (hydrolysis and assimilation by anoxic microorganisms) is still proceeding slowly and continuously.
BPs on the surface of sediments were estimated to be biodegraded more quickly than those in the water column, without forceful experimental evidence under similar conditions[30]. The properties of sediments, such as grain size and nutrient-related parameters, could affect the biodegradation of BPs placed on or buried in the sediments[28]. Grain size fractions of sediments affect the bacterial abundance, so it also indirectly affects the biodegradation rate of BPs[59], but their effect might be more or less hampered by other factors influencing the microbial community and bioactivity in natural sediments[59].
Enzymatic hydrolysis of BPs needs specific microorganisms which secrete specific extracellular enzymes (hydrolases), as well as oxidation reactions. In contrast, the microorganisms that could be involved in bioassimilation are extensive[30], so the demand for microorganisms in this step is more likely to be satisfied in the natural environment. Furthermore, inhibiting other microorganisms that are not involved in biodegradation may accelerate the biodegradation of BPs, since a broad-spectrum fungicide was found to accelerate the biodegraded rate[79]. Classically, hydrolases effective in the hydrolysis of biodegradable polyesters (PLA, PCL, and PBS) belong to cutinases (can hydrolyze ester bonds in cutin), lipases (can hydrolyze ester bonds in lipids), and polyhydroxyalkanoates (can hydrolyze ester bonds in carboxyloxoesters of bacterial polyesters); they have common features, such as extensive substrate-binding grooves for macromolecular polyesters, catalytic amino acid triad in active sites for binding subtract, and not substrate-specific[30]. Some other hydrolases that can be used for biodegradation include amidases and acetyl esterases[30].
In addition, physical shearing caused by currents/waves, wind, fauna, or sediment (especially sands) improves the fragmentation of BPs[30].
Sources of potential pollutants
Microplastics
Generation and degradation
The fragmentation of plastics in natural environments could produce secondary microplastics (MPs). BPs are supposed to produce even more MPs than conventional plastics in the short term after reaching the natural environment[80]. Then, biodegradable microplastics (BMPs) could be degraded. From another perspective, the generation of BMPs can be regarded as one of the stages of BPs’ biodegradation. However, in the natural environment, it is difficult to monitor the generation or fate of BMPs. They are even rarely detected in the open environment. As far as we know, the limited data show that PLA and PCL MPs were detected in sediments close to the wastewater outlet[81], seawater[82], and leachate[83]. On the one hand, it is related to the low consumption of BP products. On the other hand, it may be due to BPs’ high specific densities (which lead to short residence time in water, especially surface water) and shorter lifespan. Abiotic factors (UV radiation and mechanical forces) could produce BMPs from PBAT, PBS, PLA pellets, and PBAT/PLA bags, where the leading factor of BMPs’ generation rate is the original shape of BPs[84]. Lambert and Wagner also proved the release of BMPs and nanoplastics during weathering of PLA in water[85] as well as PBAT in seawater (with/without sediment)[33], PCL during enzymatic hydrolysis[46], and PHB in water under stimulated solar visible light[86], tap water, and drinking water[87]. In all the above-mentioned studies, most of the BMPs found were smaller than 50 μm.
The factors affecting the biodegradation rate of BPs also determine the period of stable existence of BMPs; due to higher specific surface areas, BMPs are supposed to be hydrolyzed more rapidly than BPs in the same environment[30]. A study supported the faster disappearance of BMPs based on the test results in the soil (ASTM D 5988-12)[88]. However, it does not mean that BMPs pose less risk than conventional MPs in the marine environment. On the one hand, a smaller size than BPs could also mean that BMPs are prone to migration on/in marine sediments[89,90] and even to depths[91] with low microbial activity. It might affect the rate of enzymatic hydrolysis and the existence time of BMPs in the marine environment. Similarly, BMPs formed in other environments could enter the marine environment [Figure 3]. Due to the promotion of UV light on the fragmentation of BPs[80], the period of BP fragmentation in soil or fresh water might be shorter than that in the marine environment. Therefore, a large proportion of secondary BMPs in the marine environment are supposed to come from other adjacent environments. For BMPs (or nanoplastics) in the marine environment, the exposure time and average abundance during exposure might be key factors for assessing their potential risk. There is no such experimental data at present, but BMPs took tens (polybutylene sebacate)[92] to hundreds (Mater-Bi)[93] of days to be mineralized in soil, and the biodegradation in the marine environment might take longer. On the other hand, the rapid degradation might pose even higher risks: the toxicity of degradation products is supposed to be stronger than the BMPs that produce them[94].
Figure 3. The generation and migration of marine biodegradable microplastics (BMPs).
Ecological risks
Among the studies involving BMPs, those concerning ecological risks are the most common [Table 4]. In view of the fact that there are only studies on PLA MPs in the marine environment, the effects of various BMPs on organisms in different environments are also listed in Table 4. It should be noted that most of these results were not obtained when exposed to the environmental abundance of BMPs.
Ecotoxicological studies related to BPs in the marine environment, freshwater, and soil
| Environment | BMPs | Abundance | Organism involved | Period (d) | Adverse effect | Effect level compared with conventional plastics | Refs. |
| Marine | PLA (~235.7 μm) | 2% of sediment weight | Arenicola marina | 31 | Respiration rate, cast production | Weaker than PVC (but PVC MPs are smaller) | [95] |
| PLA (~65.5 μm) | 80 μg/L | Ostrea edulis | 60 | Respiration rates | Similar to HDPE | [96] | |
| Scrobicularia plana | Biomasses | Similar to HDPE | |||||
| Juvenile Littorina sp. | Abundance | Similar to HDPE | |||||
| Juvenile Idotea balthica | Abundance | Similar to HDPE | |||||
| PLA (1-10 μm mostly) | 10 and 100 μg/L | Mytilus edulis | 8 | Glycerophospholipids | - | [97] | |
| PLA (~65.6 μm) | 25 μg/L | Mytilus edulis | 52 | Attachment strength, protein expression | Weaker than HDPE | [98] | |
| PLA (200-500 μm) | 100, 200, 400 mg/L | Microcosmus exasperatus | 1 | Fertilization rates | Similar to PET | [99] | |
| PLA (~65.6 μm) | 25 μg/L (Biomass of cyanobacteria, and abundance of polychaetes/ oligochaetes changed at 2.5 μg/L and 25 μg/L) | Mytilus edulis | 48 | Filtration rates | Similar to HDPE | [100] | |
| Ostrea edulis | Filtration rates | Similar to HDPE | |||||
| Biomass of cyanobacteria | Similar to HDPE | ||||||
| Abundance of Eteone picta polychaetes | Similar to HDPE | ||||||
| Abundance of Tubificoides benedii oligochaetes | Similar to HDPE | ||||||
| Freshwater | PHB (32-63 μm) | 100,000 items/individual | Gammarus fossarum | 28 | Assimilation efficiency | Weaker than PMMA | [101] |
| Wet weight change | Similar | ||||||
| PHB (~200 nm) | ~90 mg/L | Anabaena sp. PCC7120, Chlamydomonas reinhardtii | 2 | Growth | - | [86] | |
| Daphnia magna | Immobilization | - | |||||
| PLA (≤ 59 mm) | 500 mg/L (2.08 × 108 items/L) | Daphnia magna | 21 | Reproductive output | Weaker than PVC, stronger than PUR | [102] | |
| Survival | Strongest | ||||||
| PLA (~2.34 μm) | 760 and 15,020 μg/L | Physalaemus cuvieri tadpoles | 14 | Body condition indices, caudal development, triglyceride levels | - | [103] | |
| Virgin and UV irradiated PLA (5-50 μm) | 25 mg/L | Danio rerio | 11 | Skeletal development, mitochondrial dysfunction, RNA-seq of larvae | - | [104] | |
| Mater-Bi® (~41 μm) | 1 mg/L | Dreissena polymorpha | 14 | Glutathione-S-transferase activity | Weaker than PVC | [105] | |
| PLA (~2.34 μm) | 6 mg/L | Aphylla williamsoni larvae | 2 | Oxidative stress (nitrite levels, thiobarbituric acid reactive species), antioxidant activity (superoxide dismutase activity, total thiol), acetylcholinesterase activity | Stronger than PE | [106] | |
| PLA (~2.34 μm) | 3 and 9 mg/L | Danio rerio larvae | 5 | Swimming distance and speed, acetylcholinesterase activity | - | [107] | |
| Soil | Blend of PLA and PBAT (40 μm) | 1, 10 and 100 mg/L | Caenorhabditis elegans | 6 | Offspring number | Similar to LDPE | [108] |
| PLA (~120 μm) | 125, 250 and 500 g/kg | Eisenia fetida. | 2 | Avoidance test | Similar to PE | [109] | |
| 500 g/kg | 28 | Mortality | Weaker than PE | ||||
| 125, 250 and 500 g/kg | Biomass | Biomass is higher when that lower treated with PE | |||||
| 125, 250 and 500 g/kg | Offspring number | Stronger than PE | |||||
| PBAT (56.7 μm) | 20 g/kg | Arabidopsis thaliana | 49 | Rosette leaf area, plant shoot dry weight, number of fruit per plant | Stronger than LDPE | [94] | |
| 28 | Total leaf area, malondialdehyde, reactive oxygen species |
In the marine environment, Green et al. suggested that PLA MPs posed weaker effects than a conventional type of plastic (PVC) on the biological activity of a sand-dwelling lugworm[95]. A high abundance of PLA MPs (colonized by microalgae) could also elevate the respiration rates of oysters and alter assemblages (decrease species richness, total number, and body size) of benthic fauna[96]. Khalid et al. suggested that no significant effect could be found on blue mussels exposed to PLA MPs, except for the reduction in glycerophospholipids, which are important structural lipids of biological membranes[97]. Green et al. adopted a sensitive method (proteomic analysis), which found PLA MPs increased expression of a putative heavy metal binding protein and a putative detoxification enzyme of blue mussels[98]. PLA MPs could reduce the fertilization rate of solitary ascidians[99]. The response of bivalves and related habitats to the exposure of PLA MPs depend on bivalve species: oysters accelerated filtration, whereas blue mussels chose to slow down filtration; the structure of benthic community changed only in oyster-dominated habits, where the biomass of benthic cyanobacteria and numbers of polychaetes decreased, while oligochaetes increased[100].
In the freshwater environment, PHB MPs led to a weight loss of amphipods, while the biomass increased after being treated with natural or particle-free silica[101]. Nanoplastics produced from PHB MPs, instead of their leachate, induce the growth inhibition of cyanobacteria and algae and the immobilization of crustaceans[86]. Similar results (negative effects of BMPs instead of leachates) could also be found by PLA MPs on the reproductive output, body length, and mortality of crustaceans[102]. PLA MPs could inhabit tadpoles’ growth and development and reduce their lipid reserves[103]. After UV photodegradation, the efflux and detoxification rate of PLA MPs from zebrafish larvae decreased, and, accordingly, the oxidative stress effect induced mitochondrial structural damage, dysfunction, and apoptosis[104]. Mater-Bi® MPs could induce the glutathione-S-transferase activity of mussels[105], and PLA MPs increased oxidative stress, weakened antioxidant defenses, and showed anticholinergic activity in dragonfly larvae[106] and altered moving behavior and inhibited acetylcholinesterase’s activity of zebrafish larvae[107]; however, these indicators might be too sensitive as a basis for assessing toxicity.
In soil, the reproduction of nematodes could be affected by BMPs[108]. The effect of PLA MPs on the avoidance behavior, biomass, and reproduction responses of earthworms is similar to that caused by PE[109]. PBAT MPs could more severely inhibit plant growth than HDPE by disrupting the photosynthetic system and increasing the gene expression levels for drug transport[94].
In summary, most test results support the view that a high abundance of BMPs (mainly PLA MPs) could affect organisms, including biomarker contents, oxidative stress, feeding, respiration, behavior, growth, reproduction, abundance, and even mortality. It is not an inconceivable scenario in the context of the widespread use of BPs with improper solid waste management systems. However, the mode and level of these effects are highly dependent on the size and exposure abundance/period/environment of BMPs. Based on limited studies, it is not easy to draw further conclusions.
As vectors
It might still be too early to assess the physical/chemical risk of a specific type of BMP on marine organisms with few available data[87]. However, another fact to note is that BMPs could sorb toxic pollutants, including PAHs, PFASs, pesticides, pharmaceuticals, and heavy metals, and act as vectors[110]. Regarding nonpolar organic chemicals, the sorption capacity of a PAH (phenanthrene) on PBAT MPs is proven to be higher than PE and PS due to a higher proportion of rubbery subfraction, and even higher than some carbonaceous geosorbents (black carbon)[111]. The higher (than PS pellets with the larger specific surface area) sorption affinity between BMPs (PBS and PCL) and PAHs (phenanthrene/pyrene) confirmed the effect of rubbery domains; hydrogen bonding between BMPs and the polar derivatives of the above two PAHs also resulted in strong affinity[112]. The experiment concerning four chlorophenols (4-chlorophenol, 2,4-dichlorophenol, 2,4,6-trichlorophenol, and pentachlorophenol) and PLA found no significant differences between PLA and conventional MPs[113]. Regarding polar organic chemicals, the studies on fipronil, tetracycline, and ciprofloxacin supported a higher affinity with BMPs than conventional MPs due to the presence of oxygen-containing functional groups[110]. PLA MPs and conventional MPs showed no significant difference in the sorption of amphiphobic PFAFs[110]. BMPs are believed to have weaker adsorption for heavy metals (than conventional MPs), but after degradation for a period of time, their adsorption capacity for heavy metals increases[110].
In addition, the presence of MPs in the marine environment is supposed to improve horizontal gene transfer between bacterial taxa[114]. The abundance of antibiotic resistance genes and virulence factors[115] in soil treated with PBS films was larger than that of conventional plastics. It seems that conventional MPs and BMPs “choose” different bacterial hosts to spread antibiotic resistance genes and virulence factors[115].
Plastic degradation products or additives
Additives
Additives could release during biodegradation, which is influenced by the biodegradability of BPs[30]. One example is the release of carbon black[116], which maintains the mechanical properties of mulching films[30] or is used as a pigment or UV absorber[117]. Sometimes, the toxicity of plastics depends on the composition and abundance of additives, which causes the different toxicities of different products (with different additives) produced with similar plastic material (such as LDPE, PS, and PP)[118]. Moreover, PVC and PUR (which are known to contain a large amount of additives) are considered to be the most toxic with respect to bioluminescence inhibition; meanwhile, all PLA products induced similar bioluminescence inhibition with PVC and PUR in both efficiency and effect level[118], which might also be due to additives. A further study showed that the detected chemical features (possibly related to the numbers of compounds) in the extracts of BPs ranged from 614 (a PHA product) to 20,965 (a starch-based product), and it is suggested that extracts from cellulose- and starch-based BPs posed a strong in vitro toxicity[119]. It might indicate that BPs contain compounds (including additives) of the same order of magnitude with chemical features, and BPs with the highest chemical features (numbers of compounds) are the most toxic. Another study suggested that additives providing antimicrobial activity could be potentially hazardous[13]. Meanwhile, some unclear compositions of additives for proprietary reasons[30] might also pose potential risks. For example, the toxicity of MP extracts (from a PLA product) to worms was attributed to undeclared ingredients coumarin and iodocarb (presumably adsorbed) of PLA[120].
Degradation products
The degradation products of BPs could also be harmful. Some polyesters (sometimes blended with starch-based BPs), such as PBA[121], PBAT[94], and PLA[122], could release degradation products (e.g., 1,6-dioxacyclododecane-7,12-dione) and related monomers (adipic acid, lactic acid, terephthalic acid, 1,4-butanediol, etc.) to the aquatic environment; these leachates might increase proline (a plant stress marker), reduce plant development, inhibit plant growth[123], and lead to plant developmental abnormalities[121] in soil. Similar effects could occur in marine plants, such as coastal dune vegetation[124]. Moreover, adipic acid is also reported to be slightly to moderately toxic to aquatic organisms[122]. In addition to terephthalic acid and butanediol, these products were supposed to be more toxic than parent polymers such as PBAT[94].
Pesticides and other exogenous pollutants
Beriot et al. found that the average sorption percentage of 17 insecticides, 15 fungicides, and 6 herbicides on LDPE and LDPE containing pro-oxidant films was ~23%, whereas that on BP films was ~50%; at the same time, the adsorption of plastic film makes these pollutants difficult to decay in the environment[125].
Some other hazardous substances might be introduced during the production of BPs, such as endotoxins introduced in PHBV produced by Gram-negative bacteria[13].
Other impacts on the environment
Similar effects to conventional plastics
In a period after entering the natural environment, BPs may pose similar effects on habitats as conventional plastics. It is proven that corn starch-based BP films in the intertidal zone could create anoxic conditions beneath them, reduce primary productivity and organic matter in sediments, and decrease infaunal invertebrate populations, similar to those effects caused by conventional plastics[126].
The risk of entanglement of animals might be influenced by the physical and mechanical properties of BPs[30], where their degradability affects the exposure time and exposure probability indirectly.
Additional effects
The degradation time scale of BPs in the aquatic environment is designed to be shorter than that of conventional plastics. Nevertheless, it also makes the presence of BPs in nature easier to affect the ecological equilibrium and material cycle. At first, microbial communities on the surface of BPs were reported to be different from those on conventional plastics and those in the surrounding environment[24], possibly due to the surface attachment capabilities of microorganisms[127] at the beginning of attachment and the ability to utilize BPs as carbon and energy source[24] thereafter. For example, during the early stage of plastic surface biofilm formation in the marine environment, the microbial composition of biofilms on conventional plastics and aliphatic polyesters is nearly independent of plastic type[127], but a significant difference in diatom community on the surface of starch-based BP and PE was found[128]. Generally speaking, there is more biomass on the surface of BPs than on conventional plastics[24]. In the marine environment, higher cell counts and activity of bacteria could be found on PHBV than on PE, with a unique microbial community structure (hydrocarbonoclastic bacteria enriched on the plastic surface)[129]. There are differences in specific microbial communities. Pinnell and Turner compared the microbial community composition in benthic seawater of a coastal lagoon and on the surface of ceramic, PET, and PHA: the microorganism compositions on the surface of the latter three are different from that in seawater, and significantly different compositions on PHA surface were found with those of ceramic and PET, dominated by sulfate reducers[62]. Kirstein et al. suggested that marine biofilm core prokaryotic communities on the plastic substrate (including PLA) were significantly different from that on the glass substrate, but the difference in eukaryotic communities was not significant[130]. Although it is believed that the biodiversity on BP surface is lower than that on conventional plastic surfaces[24], whether this conclusion is valid in the marine environment remains to be verified. The different microbial compositions on plastic surfaces could be a hotbed of opportunistic bacteria (pathogen). For example, Vibrionaceae and Pseudoalteromonadaceae, which are rarely found in seawater and marine sediment, were detected on the surfaces of plastic debris in the seabed[131]. However, considering the difference in microbial compositions on the surface of BPs and conventional plastics, a similar conclusion cannot be extended to BPs directly. The enrichment of biomass on plastic surfaces increases the transfer rate of antibiotic resistance genes, which makes plastics hotspots for acquiring and spreading antibiotic resistance[132]. BPs and conventional plastics have different capabilities in the enrichment of antibiotic resistance genes, and this difference varies with environmental conditions; the differences in the marine environment remain to be explored[24].
Secondly, the degradation of BPs could change the surrounding environment. BPs could consume more oxygen than PE in the seabed[30]. The release of PLA MPs into sediment increases bacterial diversity (alpha diversity) and promotes nitrification and denitrification[133]; PLA MPs in lake water could alter the phytoplankton composition (eliminated cryptophytes and increased chrysophytes)[134]. A vegetable oil- and starch-based BP bag could even promote root expansion and increase the vegetative recruitment of seagrass (C. nodosa)[135]. All of the above results support the idea that a large amount of BP deposited on the seafloor can change benthic biogeochemical activities. Considering the high material density of BPs and the current production of disposable plastic products, it would not be an unimaginable scenario in the seabed of some estuaries or nearshore in the near future.
At the macro level, the widespread use of BPs could indirectly affect the environment. In the marine environment, BPs being mineralized is equivalent to being “burned”, releasing CO2 or CH4 slowly; although the whole biodegradation process may take a year or more rather than a few minutes, it does not differ much in carbon emissions calculated by year or decade. Therefore, the substitution of BPs for conventional plastics, in a sense, is to accelerate the carbon emissions from the carbon pool to the atmosphere. Note that BPs are not necessarily bio-based, and the generation of fossil minerals from carbon dioxide is much longer than the mineralization of fossil-based BPs. It is a waste of fossil minerals and accelerates the carbon cycle to use fossil-based BPs. Switching from conventional plastics to bio-based BPs is estimated to cut greenhouse gas emissions[136], but the fate of carbon fixed in plastics and its role in the greenhouse effect are still hard to assess[11]. BPs also have a greater impact than conventional plastics on issues including eutrophication and acidification[11]. From another perspective, the mass production of bio-based raw materials requires a lot of fertilizers, pesticides, cultivated land, and water resources[137]. It may not only cause additional environmental pollution but also occupy the resources required for food production. Some may argue that non-edible parts of crops, such as straw, can still be used as raw materials for BPs. However, the fact is that, by now, most straws will continue to circulate in the agricultural material flow as fertilizer and feed after harvest. Transferring them out of the agricultural sector will greatly increase the opportunity cost of agriculture. Will this demand for food and energy be transferred to marine fishing, mariculture, or offshore oil and combustible ice mining, thereby affecting the marine environment? In the context of the global food and energy shortages caused by conflict this year, it remains a possible scenario.
Limitations of the current research
At present, most studies on BPs are carried out in soil or freshwater environment. However, there are still many questions about studying BPs in the marine environment: What is the fate of BPs and BMPs in the marine environment, especially in the deep-sea environment? How different is the degradation of various BP materials in the marine environment? Will BPs and BMPs accumulate in the marine environment, and how do they migrate? How do we trace and evaluate the risk of BPs/BMPs in the marine environment?
CONCLUSION
The previous sections summarize various potential adverse impacts and hazards of BPs on the marine environment and human society, including the release of microplastics, degradation products, additives, and other pollutants, which could entangle animals or alter the material cycles of conventional plastics before decomposition, alter microbial communities on them, and affect biogeochemical activities in the adjacent environment due to their biodegradability. However, one premise of these adverse effects, namely, the large-scale use and improper abandonment of BPs, is not an established fact. Increasing the recycling rate of some plastic products and using durable materials such as metal and ceramic can reduce the proportion of disposable plastics, thus reducing the potential demand for BPs. Additionally, it should be noted that some supposed adverse effects have not been fully confirmed, since studies focusing on ecological risks of BPs in the marine environment are still limited[30]. Furthermore, some of BPs’ shortcomings can actually be overcome. For example, the application of bio-based BPs could accelerate the regeneration of BPs, making the related carbon cycle more “cost-effective”. Strengthening waste management can directly reduce the impact of BPs on the marine environment. In addition, BPs have unique advantages, such as their superiority in medical applications[138] and the ideal material for fishing gear which is often lost in the marine environment[11].
Reducing plastic pollution in the marine environment is an arduous task for human society, so no “weapon” can be given up rashly. Based on the above facts, BPs can and should be used properly: (i) BPs should not be the only option to stop plastic pollution. Comprehensive measures should be taken. In most cases, a lifestyle that reduces material and energy waste should be promoted. For those essential disposable products, proper waste management is a prerequisite for the promotion of BP products. (ii) The widespread use of non-durable organic products such as those produced by BPs should be considered when planning for carbon neutrality and ecological equilibrium. The life cycle of this kind of product may change the carbon cycle and have a long-term impact on the ecosystem in which they end up (such as landfill or marine waste accumulation points). (iii) The use of BP products should be positioned accurately. The provision of “material properties required in the application” or “biodegradability after use”[30] is fraught with contradictions. Therefore, it is very important to match the appropriate application for each BP material based on specific properties, including the persistence/exposure time in the marine environment.
To provide practical information for the application of BPs, further studies are necessary. In various marine environments with significant differences, the lifespan and environmental behavior of different biodegradable plastic products need to be correctly evaluated, which is the premise of assessing the contribution of BPs as a carbon source in the ocean. The role of additives and the risks of BMPs/leachates also need more attention. Taking BMPs as an example, the differences in the marine ecotoxicological effects caused by various materials need to be tested and confirmed, especially under the environmental abundance of BMPs. Finally, their impact on the ecosystem and material cycle (including carbon emissions) needs appropriate means or models for a comprehensive assessment.
DECLARATIONS
Author’s contribution
The author contributed solely to the article.
Availability of data and materials
Not applicable.
Financial support and sponsorship
This work was supported by National Key R&D Program of China (Grant No. 2019YFD0901101).
Conflicts of interest
The author declared that there are no conflicts of interest.
Ethical approval and consent to participate
Not applicable.
Consent for publication
Not applicable.
Copyright
© The Author(s) 2022.
REFERENCES
1. Agarwal S. Biodegradable polymers: present opportunities and challenges in providing a microplastic-free environment. Macromol Chem Phys 2020;221:2000017.
2. Schnurr REJ, Alboiu V, Chaudhary M, et al. Reducing marine pollution from single-use plastics (SUPs): a review. Mar Pollut Bull 2018;137:157-71.
3. Jia MZ. Biodegradable plastics: breaking down the facts - production, composition and environmental impact; 2020. Available from: https://www.greenpeace.org/static/planet4-eastasia-stateless/84075f56-biodegradable-plastics-report.pdf [Last accessed on 19 Sep 2022].
4. Fojt J, David J, Přikryl R, Řezáčová V, Kučerík J. A critical review of the overlooked challenge of determining micro-bioplastics in soil. Sci Total Environ 2020;745:140975.
5. IUPAC. Compendium of chemical terminology, 2nd ed. (the “Gold Book”). In: McNaught AD, Wilkinson A, Chalk SJ (online version 2019-), editors. Oxford: Blackwell Scientific Publications; 1997.
6. Qin M, Chen C, Song B, et al. A review of biodegradable plastics to biodegradable microplastics: another ecological threat to soil environments? J Clean Prod 2021;312:127816.
7. European Bioplastics, nova-Institute. Bioplastics market development update 2021. Berlin; 2021. Available from: https://www.european-bioplastics.org/news/publications/#MarketData. [Last accessed on 19 Sep 2022].
8. Ashter SA. New developments. In: Ashter SA, editor. Introduction to Bioplastics engineering. Elsevier; 2016. p. 251-74. Available from: https://www.sciencedirect.com/science/article/pii/B9780323393966000105. [Last accessed on 19 Sep 2022].
9. Kjeldsen A, Price M, Lilley C, Guzniczak E, Archer I. A review of standards for biodegradable plastics. Glasgow; 2018. Available from: https://assets.publishing.service.gov.uk/government/uploads/system/uploads/attachment_data/file/817684/review-standards-for-biodegradable-plastics-IBioIC.pdf. [Last accessed on 19 Sep 2022].
10. Ki H, Ok Park O. Synthesis, characterization and biodegradability of the biodegradable aliphatic-aromatic random copolyesters. Polymer 2001;42:1849-61.
11. Van den Oever M, Molenveld K, Van der Zee M, Bos H. Biobased and biodegradable plastics - facts and figures. Wageningen: Wageningen Food & Biobased Research; 2017.
12. Rafiqah SA, Khalina A, Harmaen AS, et al. A review on properties and application of bio-based poly(butylene succinate). Polymers (Basel) 2021;13:1436.
13. Rivera-Briso AL, Serrano-Aroca Á. Poly(3-hydroxybutyrate-co-3-hydroxyvalerate): enhancement strategies for advanced applications. Polymers (Basel) 2018;10:732.
14. Chemical Retrieval on the Web. Polymer properties database. Available from: http://www.polymerdatabase.com/home.html [Last accessed on 19 Sep 2022].
15. Shah AA, Hasan F, Hameed A, Ahmed S. Biological degradation of plastics: a comprehensive review. Biotechnol Adv 2008;26:246-65.
17. Sousa FDB. Plastic and its consequences during the COVID-19 pandemic. Environ Sci Pollut Res Int 2021;28:46067-78.
18. United States Environmental Protection Agency. Containers and packaging: product-specific data. Available from: https://www.epa.gov/facts-and-figures-about-materials-waste-and-recycling/containers-and-packaging-product-specific#C&POverview [Last accessed on 19 Sep 2022].
19. United States Environmental Protection Agency. Plastics: material-specific data. Available from: https://www.epa.gov/facts-and-figures-about-materials-waste-and-recycling/plastics-material-specific-data [Last accessed on 19 Sep 2022].
20. Merran J. 2019 International Coastal Cleanup report: the beach and beyond. Washington, DC; 2019. Available from: https://oceanconservancy.org/wp-content/uploads/2019/09/Final-2019-ICC-Report.pdf [Last accessed on 19 Sep 2022].
21. Prata JC, Silva ALP, Walker TR, Duarte AC, Rocha-Santos T. COVID-19 pandemic repercussions on the use and management of plastics. Environ Sci Technol 2020;54:7760-5.
22. Akhbarizadeh R, Dobaradaran S, Nabipour I, et al. Abandoned Covid-19 personal protective equipment along the Bushehr shores, the Persian Gulf: an emerging source of secondary microplastics in coastlines. Mar Pollut Bull 2021;168:112386.
23. Filiciotto L, Rothenberg G. Biodegradable plastics: standards, policies, and impacts. ChemSusChem 2021;14:56-72.
24. Peng C, Wang J, Liu X, Wang L. Differences in the plastispheres of biodegradable and non-biodegradable plastics: a mini review. Front Microbiol 2022;13:849147.
25. Carus M. Compostable plastic bags carry goods and hopes for the future. Hürth; 2016. Available from: https://renewable-carbon.eu/news/wp-content/uploads/2016/04/16-04-05-PR-compostable-plastic-bags-in-Europe.pdf [Last accessed on 19 Sep 2022].
26. ASTM International. ASTM standards. Available from: https://www.astm.org/catalogsearch/result/index/?q=biodegradable+plastic [Last accessed on 19 Sep 2022].
27. Choe S, Kim Y, Won Y, Myung J. Bridging three gaps in biodegradable plastics: misconceptions and truths about biodegradation. Front Chem 2021;9:671750.
28. Lott C, Eich A, Makarow D, et al. Half-life of biodegradable plastics in the marine environment depends on material, habitat, and climate zone. Front Mar Sci 2021;8:662074.
29. Elagami H, Ahmadi P, Fleckenstein JH, et al. Measurement of microplastic settling velocities and implications for residence times in thermally stratified lakes. Limnology & Oceanography 2022;67:934-45.
30. SAPEA. Biodegradability of plastics in the open environment. Berlin; 2020. Available from: https://www.sapea.info/wp-content/uploads/bop-report.pdf [Last accessed on 19 Sep 2022].
31. Haider TP, Völker C, Kramm J, Landfester K, Wurm FR. Plastics of the future? Angew Chem Int Ed Engl 2019;58:50-62.
32. Wang L, Peng Y, Xu Y, et al. Earthworms’ degradable bioplastic diet of polylactic acid: easy to break down and slow to excrete. Environ Sci Technol 2022;56:5020-8.
33. Wei XF, Bohlén M, Lindblad C, Hedenqvist M, Hakonen A. Microplastics generated from a biodegradable plastic in freshwater and seawater. Water Res 2021;198:117123.
34. Lucas N, Bienaime C, Belloy C, Queneudec M, Silvestre F, Nava-Saucedo JE. Polymer biodegradation: mechanisms and estimation techniques. Chemosphere 2008;73:429-42.
35. Wang GX, Huang D, Ji JH, Völker C, Wurm FR. Seawater-degradable polymers-fighting the marine plastic pollution. Adv Sci (Weinh) 2020;8:2001121.
36. do Val Siqueira L, Arias CILF, Maniglia BC, Tadini CC. Starch-based biodegradable plastics: methods of production, challenges and future perspectives. Current Opinion in Food Science 2021;38:122-30.
37. Rai P, Mehrotra S, Priya S, Gnansounou E, Sharma SK. Recent advances in the sustainable design and applications of biodegradable polymers. Bioresour Technol 2021;325:124739.
38. European Bioplastics, nova-Institute. Bioplastics market data 2017. Berlin; 2017. Available from: https://docs.european-bioplastics.org/publications/market_data/2017/Report_Bioplastics_Market_Data_2017.pdf [Last accessed on 19 Sep 2022].
39. European Bioplastics, nova-Institute. Bioplastics market development update 2019. Berlin; 2019. Available from: http://www.european-bioplastics.org/news/publications/ [Last accessed on 19 Sep 2022].
40. European Bioplastics, nova-Institute. Bioplastics market data 2018. Berlin; 2018. Available from: https://www.european-bioplastics.org/new-market-data-the-positive-trend-for-the-bioplastics-industry-remains-stable/ [Last accessed on 19 Sep 2022].
41. European Bioplastics, nova-Institute. Bioplastic market development update 2020. Berlin; 2020. Available from: https://docs.european-bioplastics.org/conference/Report_Bioplastics_Market_Data_2020_short_version.pdf [Last accessed on 19 Sep 2022].
42. Min K, Cuiffi JD, Mathers RT. Ranking environmental degradation trends of plastic marine debris based on physical properties and molecular structure. Nat Commun 2020;11:727.
43. Muhamad II, Joon LK, Noor MAM. Comparing the degradation of poly-β-(hydroxybutyrate), Poly-β-(hydroxybutyrate-co-valerate)(PHBV) and PHBV/cellulose triacetate blend. Malaysian Polym J 2006;1:39-46.
44. Meereboer KW, Misra M, Mohanty AK. Review of recent advances in the biodegradability of polyhydroxyalkanoate (PHA) bioplastics and their composites. Green Chem 2020;22:5519-58.
45. Burkersroda FV, Schedl L, Göpferich A. Why degradable polymers undergo surface erosion or bulk erosion. Biomaterials 2002;23:4221-31.
46. Wei XF, Capezza AJ, Cui Y, et al. Millions of microplastics released from a biodegradable polymer during biodegradation/enzymatic hydrolysis. Water Res 2022;211:118068.
47. Volova T, Boyandin A, Vasiliev A, et al. Biodegradation of polyhydroxyalkanoates (PHAs) in tropical coastal waters and identification of PHA-degrading bacteria. Polym Degrad Stab 2010;95:2350-9.
48. Bagheri AR, Laforsch C, Greiner A, Agarwal S. Fate of so-called biodegradable polymers in seawater and freshwater. Glob Chall 2017;1:1700048.
49. Catarci Carteny C, Blust R. Not only diamonds are forever: degradation of plastic films in a simulated marine environment. Front Environ Sci 2021;9:662844.
50. Tsuji H, Suzuyoshi K. Environmental degradation of biodegradable polyesters 1. Poly(ε-caprolactone), poly[(R)-3-hydroxybutyrate], and poly(L-lactide) films in controlled static seawater. Polym Degrad Stab 2002;75:347-55.
51. Narancic T, Verstichel S, Reddy Chaganti S, et al. Biodegradable plastic blends create new possibilities for end-of-life management of plastics but they are not a panacea for plastic pollution. Environ Sci Technol 2018;52:10441-52.
52. Wang G, Huang D, Zhang W, Ji J. Degradation performance of typical biodegradable polyesters in seawater. J Funct Polym 2020;33:492-9.
53. Delacuvellerie A, Benali S, Cyriaque V, et al. Microbial biofilm composition and polymer degradation of compostable and non-compostable plastics immersed in the marine environment. J Hazard Mater 2021;419:126526.
54. Cocca M, Falco F De, Gentile G, Avolio R, Errico ME, et al. Degradation of biodegradable plastic buried in sand. In: Cocca M, Di Pace E, Errico ME, et al., editors. Proceedings of the International Conference on Microplastic Pollution in the Mediterranean Sea. Switzerland: Springer International Publishing; 2018. p. 205-10. Available from: http://link.springer.com/10.1007/978-3-319-71279-6 [Last accessed on 19 Sep 2022].
55. Beltrán-Sanahuja A, Casado-Coy N, Simó-Cabrera L, Sanz-Lázaro C. Monitoring polymer degradation under different conditions in the marine environment. Environ Pollut 2020;259:113836.
56. Tachibana K, Urano Y, Numata K. Biodegradability of nylon 4 film in a marine environment. Polym Degrad Stab 2013;98:1847-51.
57. Kasuya K, Takagi K, Ishiwatari S, Yoshida Y, Doi Y. Biodegradabilities of various aliphatic polyesters in natural waters. Polym Degrad Stab 1998;59:327-32.
58. Thellen C, Coyne M, Froio D, Auerbach M, Wirsen C, Ratto JA. A processing, characterization and marine biodegradation study of melt-extruded polyhydroxyalkanoate (PHA) films. J Polym Environ 2008;16:1-11.
59. Eich A, Weber M, Lott C. Disintegration half-life of biodegradable plastic films on different marine beach sediments. PeerJ 2021;9:e11981.
60. Accinelli C, Saccà ML, Mencarelli M, Vicari A. Deterioration of bioplastic carrier bags in the environment and assessment of a new recycling alternative. Chemosphere 2012;89:136-43.
62. Pinnell LJ, Turner JW. Shotgun metagenomics reveals the benthic microbial community response to plastic and bioplastic in a coastal marine environment. Front Microbiol 2019;10:1252.
63. Desrousseaux C, Sautou V, Descamps S, Traoré O. Modification of the surfaces of medical devices to prevent microbial adhesion and biofilm formation. J Hosp Infect 2013;85:87-93.
64. Laycock B, Nikolić M, Colwell JM, et al. Lifetime prediction of biodegradable polymers. Progress in Polymer Science 2017;71:144-89.
65. Teramoto N, Urata K, Ozawa K, Shibata M. Biodegradation of aliphatic polyester composites reinforced by abaca fiber. Prog Polym Sci 2004;86:401-9.
66. Díaz A, Katsarava R, Puiggalí J. Synthesis, properties and applications of biodegradable polymers derived from diols and dicarboxylic acids: from polyesters to poly(ester amide)s. Int J Mol Sci 2014;15:7064-123.
67. Jacquin J, Callac N, Cheng J, et al. Microbial diversity and activity during the biodegradation in seawater of various substitutes to conventional plastic cotton swab sticks. Front Microbiol 2021;12:604395.
68. Okoye CO, Addey CI, Oderinde O, et al. Toxic chemicals and persistent organic pollutants associated with micro-and nanoplastics pollution. Chem Eng J Adv 2022;11:100310.
69. Woolnough CA, Yee LH, Charlton TS, Foster LJ. A tuneable switch for controlling environmental degradation of bioplastics: addition of isothiazolinone to polyhydroxyalkanoates. PLoS One 2013;8:e75817.
70. Chen H. Assessment of biodegradation in different environmental compartments of blends and composites based on microbial poly(hydroxyalkanoate)s. University of Pisa; 2012. Available from: https://etd.adm.unipi.it/theses/available/etd-02222013-142550/unrestricted/Thesis__Chen_Haiyan.pdf [Last accessed on 19 Sep 2022].
71. Gerritse J, Leslie HA, de Tender CA, Devriese LI, Vethaak AD. Fragmentation of plastic objects in a laboratory seawater microcosm. Sci Rep 2020;10:10945.
72. Lambert S, Wagner M. Environmental performance of bio-based and biodegradable plastics: the road ahead. Chem Soc Rev 2017;46:6855-71.
73. De Hoe GX, Zumstein MT, Getzinger GJ, et al. Photochemical transformation of poly(butylene adipate- co-terephthalate) and its effects on enzymatic hydrolyzability. Environ Sci Technol 2019;53:2472-81.
74. Lott C, Eich A, Unger B, et al. Field and mesocosm methods to test biodegradable plastic film under marine conditions. PLoS One 2020;15:e0236579.
75. Amobonye A, Bhagwat P, Singh S, Pillai S. Plastic biodegradation: frontline microbes and their enzymes. Sci Total Environ 2021;759:143536.
76. Lin Z, Jin T, Zou T, et al. Current progress on plastic/microplastic degradation: fact influences and mechanism. Environ Pollut 2022;304:119159.
77. Tian C, Lv J, Zhang W, et al. Accelerated degradation of microplastics at the liquid interface of ice crystals in frozen aqueous solutions. Angew Chem Int Ed Engl 2022;61:e202206947.
78. Zhang B, Yang X, Liu L, et al. Spatial and seasonal variations in biofilm formation on microplastics in coastal waters. Sci Total Environ 2021;770:145303.
79. Li R, Liu Y, Sheng Y, Xiang Q, Zhou Y, Cizdziel JV. Effect of prothioconazole on the degradation of microplastics derived from mulching plastic film: apparent change and interaction with heavy metals in soil. Environ Pollut 2020;260:113988.
80. Yang Y, Li Z, Yan C, et al. Kinetics of microplastic generation from different types of mulch films in agricultural soil. Sci Total Environ 2022;814:152572.
81. Granberg M, von Friesen LW, Bach L, Collard F, Strand J, et al. Anthropogenic microlitter in wastewater and marine samples from Ny-Ålesund, Barentsburg and Signehamna, Svalbard. Stockholm; 2019. Available from: https://pure.au.dk/portal/files/152355861/C373march.pdf [Last accessed on 19 Sep 2022].
82. Kazour M, Jemaa S, Issa C, Khalaf G, Amara R. Microplastics pollution along the Lebanese coast (Eastern Mediterranean Basin): occurrence in surface water, sediments and biota samples. Sci Total Environ 2019;696:133933.
83. Sun J, Zhu ZR, Li WH, et al. Revisiting microplastics in landfill leachate: unnoticed tiny microplastics and their fate in treatment works. Water Res 2021;190:116784.
84. Tong H, Zhong X, Duan Z, et al. Micro- and nanoplastics released from biodegradable and conventional plastics during degradation: formation, aging factors, and toxicity. Sci Total Environ 2022;833:155275.
85. Lambert S, Wagner M. Formation of microscopic particles during the degradation of different polymers. Chemosphere 2016;161:510-7.
86. González-pleiter M, Tamayo-belda M, Pulido-reyes G, et al. Secondary nanoplastics released from a biodegradable microplastic severely impact freshwater environments. Environ Sci: Nano 2019;6:1382-92.
87. Shruti VC, Kutralam-Muniasamy G. Bioplastics: missing link in the era of microplastics. Sci Total Environ 2019;697:134139.
88. Innocenti FD. Biodegradable plastics do not form chemically persistent microplastics. In: Cocca M, Di Pace E, Errico ME, et al., editors. Proceedings of the 2nd International Conference on Microplastic Pollution in the Mediterranean Sea. Switzerland: Springer International Publishing; 2020. p. 82-8. Available from: http://dx.doi.org/10.1007/978-3-030-45909-3_15 [Last accessed on 19 Sep 2022].
89. Harris PT. The fate of microplastic in marine sedimentary environments: a review and synthesis. Mar Pollut Bull 2020;158:111398.
90. Dong Z, Qiu Y, Zhang W, Yang Z, Wei L. Size-dependent transport and retention of micron-sized plastic spheres in natural sand saturated with seawater. Water Res 2018;143:518-26.
91. Pohl F, Eggenhuisen JT, Kane IA, Clare MA. Transport and burial of microplastics in deep-marine sediments by turbidity currents. Environ Sci Technol 2020;54:4180-9.
92. Chinaglia S, Tosin M, Degli-innocenti F. Biodegradation rate of biodegradable plastics at molecular level. Polym Degrad Stab 2018;147:237-44.
93. Tosin M, Pischedda A, Degli-innocenti F. Biodegradation kinetics in soil of a multi-constituent biodegradable plastic. Polym Degrad Stab 2019;166:213-8.
94. Liu J, Wang P, Wang Y, et al. Negative effects of poly(butylene adipate-co-terephthalate) microplastics on Arabidopsis and its root-associated microbiome. J Hazard Mater 2022;437:129294.
95. Green DS, Boots B, Sigwart J, Jiang S, Rocha C. Effects of conventional and biodegradable microplastics on a marine ecosystem engineer (Arenicola marina) and sediment nutrient cycling. Environ Pollut 2016;208:426-34.
96. Green DS. Effects of microplastics on European flat oysters, Ostrea edulis and their associated benthic communities. Environ Pollut 2016;216:95-103.
97. Khalid A, Zalouk-Vergnoux A, Benali S, et al. Are bio-based and biodegradable microplastics impacting for blue mussel (Mytilus edulis)? Mar Pollut Bull 2021;167:112295.
98. Green DS, Colgan TJ, Thompson RC, Carolan JC. Exposure to microplastics reduces attachment strength and alters the haemolymph proteome of blue mussels (Mytilus edulis). Environ Pollut 2019;246:423-34.
99. Anderson G, Shenkar N. Potential effects of biodegradable single-use items in the sea: Polylactic acid (PLA) and solitary ascidians. Environ Pollut 2021;268:115364.
100. Green DS, Boots B, O’Connor NE, Thompson R. Microplastics affect the ecological functioning of an important biogenic habitat. Environ Sci Technol 2017;51:68-77.
101. Straub S, Hirsch PE, Burkhardt-Holm P. Biodegradable and petroleum-based microplastics do not differ in their ingestion and excretion but in their biological effects in a freshwater invertebrate gammarus fossarum. Int J Environ Res Public Health 2017;14:774.
102. Zimmermann L, Göttlich S, Oehlmann J, Wagner M, Völker C. What are the drivers of microplastic toxicity? Environ Pollut 2020;267:115392.
103. Malafaia G, Nascimento ÍF, Estrela FN, et al. Green toxicology approach involving polylactic acid biomicroplastics and neotropical tadpoles: (Eco)toxicological safety or environmental hazard? Sci Total Environ 2021;783:146994.
104. Zhang X, Xia M, Su X, et al. Photolytic degradation elevated the toxicity of polylactic acid microplastics to developing zebrafish by triggering mitochondrial dysfunction and apoptosis. J Hazard Mater 2021;413:125321.
105. Magni S, Bonasoro F, Della Torre C, Parenti CC, Maggioni D, Binelli A. Plastics and biodegradable plastics: ecotoxicity comparison between polyvinylchloride and Mater-Bi® micro-debris in a freshwater biological model. Sci Total Environ 2020;720:137602.
106. Chagas TQ, Araújo APDC, Malafaia G. Biomicroplastics versus conventional microplastics: an insight on the toxicity of these polymers in dragonfly larvae. Sci Total Environ 2021;761:143231.
107. Oliveira JPJ, Estrela FN, Rodrigues ASL, Guimarães ATB, Rocha TL, Malafaia G. Behavioral and biochemical consequences of Danio rerio larvae exposure to polylactic acid bioplastic. J Hazard Mater 2021;404:124152.
108. Schöpfer L, Menzel R, Schnepf U, et al. Microplastics effects on reproduction and body length of the soil-dwelling nematode caenorhabditis elegans. Front Environ Sci 2020;8:41.
109. Ding W, Li Z, Qi R, et al. Effect thresholds for the earthworm Eisenia fetida: toxicity comparison between conventional and biodegradable microplastics. Sci Total Environ 2021;781:146884.
110. Torres FG, Dioses-Salinas DC, Pizarro-Ortega CI, De-la-Torre GE. Sorption of chemical contaminants on degradable and non-degradable microplastics: recent progress and research trends. Sci Total Environ 2021;757:143875.
111. Zuo LZ, Li HX, Lin L, et al. Sorption and desorption of phenanthrene on biodegradable poly(butylene adipate co-terephtalate) microplastics. Chemosphere 2019;215:25-32.
112. Zhao L, Rong L, Xu J, Lian J, Wang L, Sun H. Sorption of five organic compounds by polar and nonpolar microplastics. Chemosphere 2020;257:127206.
113. Tubić A, Lončarski M, Maletić S, et al. Significance of chlorinated phenols adsorption on plastics and bioplastics during water treatment. Water 2019;11:2358.
114. Liu Y, Liu W, Yang X, Wang J, Lin H, Yang Y. Microplastics are a hotspot for antibiotic resistance genes: progress and perspective. Sci Total Environ 2021;773:145643.
115. Song R, Sun Y, Li X, et al. Biodegradable microplastics induced the dissemination of antibiotic resistance genes and virulence factors in soil: a metagenomic perspective. Sci Total Environ 2022;828:154596.
116. Sintim HY, Bary AI, Hayes DG, et al. Release of micro- and nanoparticles from biodegradable plastic during in situ composting. Sci Total Environ 2019;675:686-93.
117. Long CM, Nascarella MA, Valberg PA. Carbon black vs. black carbon and other airborne materials containing elemental carbon: physical and chemical distinctions. Environ Pollut 2013;181:271-86.
118. Zimmermann L, Dierkes G, Ternes TA, Völker C, Wagner M. Benchmarking the in vitro toxicity and chemical composition of plastic consumer products. Environ Sci Technol 2019;53:11467-77.
119. Zimmermann L, Dombrowski A, Völker C, Wagner M. Are bioplastics and plant-based materials safer than conventional plastics? Environ Int 2020;145:106066.
120. Klein K, Piana T, Lauschke T, et al. Chemicals associated with biodegradable microplastic drive the toxicity to the freshwater oligochaete Lumbriculus variegatus. Aquat Toxicol 2021;231:105723.
121. Balestri E, Menicagli V, Ligorini V, Fulignati S, Raspolli Galletti AM, Lardicci C. Phytotoxicity assessment of conventional and biodegradable plastic bags using seed germination test. Ecological Indicators 2019;102:569-80.
123. Serrano-ruíz H, Eras J, Martín-closas L, Pelacho A. Compounds released from unused biodegradable mulch materials after contact with water. Polym Degrad Stab 2020;178:109202.
124. Menicagli V, Balestri E, Lardicci C. Exposure of coastal dune vegetation to plastic bag leachates: a neglected impact of plastic litter. Sci Total Environ 2019;683:737-48.
125. Beriot N, Zomer P, Zornoza R, Geissen V. A laboratory comparison of the interactions between three plastic mulch types and 38 active substances found in pesticides. PeerJ 2020;8:e9876.
126. Green DS, Boots B, Blockley DJ, Rocha C, Thompson R. Impacts of discarded plastic bags on marine assemblages and ecosystem functioning. Environ Sci Technol 2015;49:5380-9.
127. Odobel C, Dussud C, Philip L, et al. Bacterial abundance, diversity and activity during long-term colonization of non-biodegradable and biodegradable plastics in seawater. Front Microbiol 2021;12:734782.
128. Eich A, Mildenberger T, Laforsch C, Weber M. Biofilm and diatom succession on polyethylene (PE) and biodegradable plastic bags in two marine habitats: early signs of degradation in the pelagic and benthic zone? PLoS One 2015;10:e0137201.
129. Dussud C, Hudec C, George M, et al. Colonization of non-biodegradable and biodegradable plastics by marine microorganisms. Front Microbiol 2018;9:1571.
130. Kirstein IV, Wichels A, Krohne G, Gerdts G. Mature biofilm communities on synthetic polymers in seawater - specific or general? Mar Environ Res 2018;142:147-54.
131. De Tender CA, Devriese LI, Haegeman A, Maes S, Ruttink T, Dawyndt P. Bacterial community profiling of plastic litter in the belgian part of the north sea. Environ Sci Technol 2015;49:9629-38.
132. Yang K, Chen QL, Chen ML, et al. Temporal dynamics of antibiotic resistome in the plastisphere during microbial colonization. Environ Sci Technol 2020;54:11322-32.
133. Seeley ME, Song B, Passie R, Hale RC. Microplastics affect sedimentary microbial communities and nitrogen cycling. Nat Commun 2020;11:2372.
134. Yokota K, Mehlrose M. Lake phytoplankton assemblage altered by irregularly shaped pla body wash microplastics but not by PS calibration beads. Water 2020;12:2650.
135. Balestri E, Menicagli V, Vallerini F, Lardicci C. Biodegradable plastic bags on the seafloor: a future threat for seagrass meadows? Sci Total Environ 2017;605-606:755-63.
136. Posen ID, Jaramillo P, Landis AE, Griffin WM. Greenhouse gas mitigation for U.S. plastics production: energy first, feedstocks later. Environ Res Lett 2017;12:034024.
137. Paunonen S, Kamppuri T, Katajainen L, Hohenthal C, Heikkilä P, Harlin A. Environmental impact of cellulose carbamate fibers from chemically recycled cotton. J Clean Prod 2019;222:871-81.
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Chen, H. Biodegradable plastics in the marine environment: a potential source of risk?. Water Emerg. Contam. Nanoplastics. 2022, 1, 16. http://dx.doi.org/10.20517/wecn.2022.11
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