# Challenges and opportunities in home-composting of biodegradable plastics: A comprehensive review

Jules Bellon, Feriel Bacoup, Aamir Shehzad, Richard Gattin

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Jules Bellon, Feriel Bacoup, Aamir Shehzad, Richard Gattin. Challenges and opportunities in home-composting of biodegradable plastics: A comprehensive review. Results in engineering, 2025, 28, pp.107535. 10.1016/j.rineng.2025.107535. hal-05357359

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## Review article

# Challenges and opportunities in home-composting of biodegradable plastics: A comprehensive review

Jules Bellon* ID, Feriel Bacoup, Aamir Shehzad, Richard Gattin

UniLaSalle, Transformation and Agro-Resources Research Unit (ULR 7519 UniLaSalle-Université d'Artois), 76130 Mont-Saint-Aignan, France

### ARTICLE INFO

**Keywords:**

Home-composting

Biodegradable polymers

Biodegradability assessment

Mesophilic conditions

Microplastics

Circular economy

### ABSTRACT

Biodegradable plastics offer a promising alternative to conventional plastics, but their environmental benefits depend heavily on appropriate end-of-life strategies. Home composting stands out as a decentralized and accessible option with low infrastructure needs and strong potential to involve citizens in sustainable waste management. However, its effectiveness remains uncertain due to fluctuating conditions—such as temperature and humidity—arising from the small scale of composting systems, variability in household practices, and geographic or seasonal influences. These constraints pose significant challenges for the reliable home compostability of bioplastics. This comprehensive review examines the characteristics of decentralized composting systems and their influence on the biodegradation of biodegradable plastics. A total of 48 experimental studies were analysed: 27 conducted under controlled laboratory conditions and 21 under real-life or uncontrolled environments. These studies assessed biodegradability using respirometric techniques, weight loss measurements, or both. Three major factors were identified as key drivers of biodegradation rates: (i) the composition of the plastic, including polymer type, chemical structure, and the presence of fillers or additives; (ii) composting system parameters, such as aeration, volume, and temperature stability; and (iii) the properties of the composting medium, including microbial community composition and nutrient availability. In addition, the review discusses the risk of microplastic persistence, and the methodological heterogeneity observed across studies, which complicates the assessment and comparison of results. Practical recommendations to enhance home composting performance are provided throughout the paper, supporting the development of robust, context-sensitive home-compostability assessment frameworks.

## 1. Introduction

The global production of petroleum-based plastics has been growing at an exponential rate since 1950, due to the widespread use of these plastics in numerous daily items, and the multiplication of manufactured products, leading to abundant pollution within all biospheric sections [1–3]. The pressure exerted by these materials on the environment is explained, on the one hand, by their dispersion and accumulation within natural environments, and is due to their low degradation rates, thus disrupting the functioning of ecosystems [4]. On the other hand, the production of petroleum-based plastics contributes to several harmful phenomena, beyond their environmental accumulation. In particular, it plays a notable role in global warming by releasing substantial amounts of greenhouse gases (GHG), accounting for approximately 4 % of global CO₂-equivalent emissions [5]. The packaging sector, particularly food, is the main consumer of plastic (39.9 % in 2015, EU) [4]. Reducing the use

of plastics remains the preferred way to overcome many problems arising from their use. However, a plastic-free world seems difficult to conceive, by virtue of the considerable impacts which their technological properties and low cost confer on our consumption patterns [6,7].

Single-use plastics currently constitute the bulk of the packaging market, due to their excellent thermal and mechanical properties [8,9]. These include a diverse range of common consumer products, which are discarded after a single use. These are not intended to be reused during their life cycle for an identical purpose for which these were designed, and are still slightly recycled (27 %, France 2018; 5–6 %, USA 2021). These are therefore highly susceptible to becoming waste [10,11] which is most of the time buried (78–80 %, USA 2021) or incinerated (15–16 %, USA 2021) [12]. These two end-of-life management pathways (i.e. landfill and incineration) greatly increase the environmental impact of these materials: air pollution, groundwater contamination, reduced soil quality, proliferation of plastic nano-microparticles [13–15].

<footnote>* Corresponding author.</footnote>

<footnote>E-mail address: jules.bellon@unilasalle.fr (J. Bellon).</footnote>
Some high molecular weight polymers, including aliphatic polyesters (e.g. PLA, PBS, PHB) and aliphatic-aromatic copolymers (e.g. PBAT), can undergo biodegradation within a relatively short period. This is primarily due to the presence of chemically labile bonds, such as ester linkages, which are readily cleaved under certain environmental conditions. For a given biodegradable polymer, this process is governed by the interplay between environmental (i.e. abiotic) conditions and biological (i.e. biotic) activity. Its kinetics are strongly influenced by a third key factor: the intrinsic properties of the polymer itself, such as crystallinity and molecular weight [16,17].

Biodegradation involves a reduction in the molecular weight of these polymers, and the transport of low-molecular-weight compounds across the cell membranes of microorganisms; these compounds are then metabolized. This final step is referred to as bio-assimilation. Biodegradation should be distinguished from disintegration, which strictly refers to the physical breakdown of a material into smaller fragments [18]. Disintegration is typically assessed through measurements of weight loss or surface erosion. In contrast, biodegradation is evaluated using more complex methods, which involve monitoring the temporal release of gaseous chemical species (e.g. CO₂, CH₄) produced by microbial metabolic activity during the breakdown of the material [19,20].

Composting represents a particularly relevant organic waste recovery pathway, as it enables the controlled transformation of organic matter into stabilized compost, rich in nutrients and suitable for reintegration into soils [21]. By promoting the return of organic matter to the soil and reducing reliance on mineral fertilizers, this sector fits squarely within a circular economy framework [22,23]. In this context, the integration of biodegradable plastics compatible with composting conditions appears to be a promising approach to improving the management of certain single-use plastic wastes. Composting systems can be broadly categorized into two main types: centralized (or industrial) composting, and decentralized systems, encompassing both home-composting and community-scale initiatives. This distinction reflects differences in scale, operational control, and waste input characteristics [24,25].

Centralized composting has several limitations: conveyed routes, additional sorting step requiring the use of sometimes expensive separation methods (IR laser), construction and maintenance of dedicated infrastructure, training and use of specific personnel [26,27]. Although it is more efficient due to its ability to reach a real phase qualified as thermophilic, involving a higher temperature rise (~55–65 °C), for a sufficient time, compared to its decentralized counterpart (i.e. domestic and community composting).

The biodegradative capacity of decentralized composting is lower compared to centralized composting. This fact is mainly explained by the absence of a real thermophilic phase during the process. The lower volumes (< 2.5 m³) of decomposing organic matter limit the temperature rise within the compost pile [28]. Smaller volumes of decomposing matter result in a larger surface-to-volume ratio, and consequently preeminent heat loss, through thermal convection [29].

However, as opposed to centralized composting, the large-scale deployment and logistical implementation of the decentralized route is simpler: it does not require the construction of dedicated structures, the training of qualified personnel, and the implementation of waste collection [30]. The compost obtained has a very high agronomic quality, characterized by the low presence of contaminants, a low level of heavy metals, and a high concentration of nutrients for the higher plants [31,32]. The co-location of biowaste production and disposal areas drastically reduces the route taken by biowaste, thus reducing the consumption of fossil fuels associated with their transport [33]. Active participation of the people in the management of organic waste results in their heightened consciousness of prevailing environmental concerns [34].

For these reasons, home composting stands out as a relevant and promising end-of-life management option for biodegradable plastics. Nevertheless, it remains a challenge for most polymers currently

available on the market, and even more so for the manufactured plastic products derived from them. This limitation is mainly due to their lower biodegradability under the mesophilic conditions typically found in home composting systems [35]. This inherently restricts the range of plastic formulations suitable for this end-of-life management route, thereby hindering the broader development of organic recycling pathways for plastic waste.

So far, few studies have focused on the home composting of biodegradable plastics, in contrast to the more extensively studied industrial composting. This gap limits our understanding of the key factors influencing the biodegradability of major biodegradable polymers under home composting conditions. Moreover, unlike industrial composting, home composting is characterized by more variable environmental conditions (e.g. temperature, humidity, microbial diversity and activity), primarily due to differences in individual management practices. In addition, the smaller volume of material involved makes the composting environment more susceptible to external factors such as geographical location, season, and the placement of the composting unit [36–39].

In this literature review, we focus specifically on the home composting of biodegradable plastics. First, we identify the key characteristics that distinguish home composting from industrial composting and examine the implications of these differences for the biodegradation of plastics and the relevant specification standards. We then present an overview of the main biodegradable polymers that are likely to follow this end-of-life management route, along with plastics marketed as biodegradable.

Next, we provide a comprehensive overview of published studies that have tested the biodegradability of plastics under mesophilic composting conditions. This allows us to compile an inventory of tested materials and to identify key sources of variability.

We also offer a critical analysis of the methodologies used in current specification standards related to the home compostability of biodegradable plastics. Additionally, throughout the review, we provide practical recommendations aimed at improving the home compostability of biodegradable plastics, as well as research perspectives to guide future academic investigations in this field.

Finally, we explore the environmental impact of biodegradable plastics in home composting systems, with a particular focus on the potential effects of nano- and microplastics on soil organisms and overall soil ecosystem functioning.

## 2. Plastic biodegradability: from molecular structure to environmental fate

Plastics include a wide range of materials composed of at least one high-molecular-weight organic polymer. To impart plastics specific properties and/or reduce their cost, these typically incorporate additional substances known as additives. These are most often plasticizers, stabilizers, fillers, reinforcing agents, flame retardants and colorants [40].

Polymers can be divided into four categories: (i) non-biodegradable (i.e. conventional) petro-based polymers; (ii) biodegradable petro-based polymers; (iii) non-biodegradable biobased polymers; (iv) biodegradable biobased polymers. The biodegradability of a polymer material is an extrinsic property. Since it depends on two components: the intrinsic properties of the material, and the environment in which it degrades (i.e. biotic and abiotic) [41].

The biodegradation process takes place in two successive stages. (i) The breakdown of macromolecular chains, leading to the formation of lower molecular weight chains. (ii) Bio-assimilation of low-molecular-weight compounds by microorganisms, i.e. transport across the cell membrane, and subsequent metabolism (i.e. catabolic and anabolic reactions). The second stage commences when the molecular weight of compounds declines below a critical value [41].

The main mechanisms involved in the initial phase of the degradation process include: (i) non-enzymatic chemical hydrolysis, (ii)
enzymatic hydrolysis, (iii) oxidation, and (iv) photo-oxidation [42,43].

In most cases, completion of the first stage determines the rate of biodegradation for a polymer-environment pair. It can be seen as a bottleneck in the biodegradation process. If it is too slow, then the polymer persists in the environment and its accumulation leads to deleterious effects on ecosystems [44].

The presence of easily cleavable chemical bonds (by hydrolysis or oxidation) along the carbon chain is the most important factor in achieving the first step. Typically, biodegradable polymers most often possess one or more of the following hydrolysable chemical bonds: esters (R-COO-R'), amides (R-CONH2) or glycosidic (R—O-R') [45]. Most conventional petroleum-based polymers lack such chemical bonds, which explains their persistence. This is particularly true of PE, PP, PS and PVC. However, some conventional petroleum-based polymers have these chemical bonds, but are not considered biodegradable. This is particularly true for PET (i.e. polyester). In this case, the slow progress of the first stage (i.e. decomposition of macromolecular chains) is explained by the high intensity with which the chains interact with each other, which modulates the propensity of hydrolysable bonds to react and break [46].

The biodegradability of so-called biodegradable plastics strongly depends on the environment in which the process takes place. Key factors include physico-chemical parameters (e.g. temperature, humidity) and biological conditions (e.g. microbial activity and diversity) [47]. It is therefore important to associate the term *biodegradable plastic* with the specific environments in which these plastics can biodegrade. This clarification would help prevent misunderstandings among consumers regarding the conditions under which these materials can truly break down.

Under natural conditions, their biodegradability is generally limited. Therefore, although these materials are considered less persistent in environments with low biodegradation capacity—such as aquatic systems—compared to conventional plastics, they still require appropriate end-of-life management. Without it, their accumulation in the environment after use remains a concern [48].

These end-of-life management pathways can be ranked according to their biodegradation potential for biodegradable plastics, in decreasing order as follows: centralized composting > anaerobic digestion > decentralized composting > soil > freshwater sludge >> marine environments > landfill [7,49].

Abundance, diversity and microbial activity are essential biological elements making it possible to discriminate environments according to their biodegradative capacity. On the abiotic level, the temperature parameter, not without being the only determinant, is an essential component [50].

While composting and anaerobic digestion offer greater biodegradation potential than decentralized composting (i.e. home and community composting), they are both centralized end-of-life management pathways. Unlike home composting, they require more complex logistics and tend to diverge from the principles of a circular economy [51–53].

# 3. Characteristics of home composting

## 3.1. Centralized and decentralized composting

Two types of composting systems can be distinguished based on how different sources of emissions from biowaste are bundled together. Differentiating amounts of waste are handled separately as a result. Centralized composting, also known as industrial, takes place on dedicated platforms, often located on the outskirts of large urban areas. It involves the centralization of waste through collection, and consequently its transport by road [54], and allows the unit treatment of large quantities of biowaste (1–5 tons for tank composting, and > 10 tons for windrow composting) [55]. Decentralized composting does not entail the concentration of bio-waste. Instead, the treatment facilities are

situated near the emission zones and handle only smaller quantities of waste. It is divided into two subcategories: (i) community composting implies the practice of collecting and treating waste at the level of a community of individuals living in a specific geographic area. The management of this sort of composting is left to the municipality, a private company, or the community itself. (ii) Home-composting is the process of treating waste at the level of a household, featuring the individual use of the compost obtained [56,57].

Home-composting must take longer (6–12 months) than industrial composting (2–3 months) to attain a stable condition for the organic matter of decomposed biowaste and is explained by the shorter thermophilic phase [50]. The compost obtained from decentralized units (domestic and community) is recognized as having a very high agro-nomic quality, characterized by low heavy metal contents, a high fertilizing value, and a satisfactory stabilization of organic matter [58, 59].

Compared to centralized treatment methods, decentralized composting is characterized by its ease of implementation, low infrastructure requirements, and its potential to actively engage citizens in the management of their biowaste [56].

In Shiraz (Iran), the introduction of a system covering 10 % of organic waste enabled the municipality to reduce overall waste management costs by approximately 5 % over a 10-year period, while cutting CO2-equivalent emissions by >19,000 tonnes per year [51].

Similar findings were reported in Poland, where a municipal incentive program near Warsaw showed that home composting can be economically beneficial for communities. Beyond a certain participation threshold, savings from reduced collection, transport, and centralized treatment of biowaste—initially landfilled—can offset initial investments, such as subsidized composters. Local authorities supported the initiative through financial aid, reduced waste collection fees for participants, and awareness campaigns [52].

These examples suggest that home composting can be a cost-effective and environmentally sound alternative to centralized systems in certain contexts. By contrast, industrial composting facilities offer much greater treatment capacity and better control of the physico-chemical conditions favourable to biodegradation, including biodegradable plastics. However, such performance comes at the cost of significantly higher infrastructure investment and operating expenses. These facilities require advanced thermal and aeration control systems, odour management infrastructure, strict regulatory compliance mechanisms, and trained personnel [60].

A comparative study conducted in Québec clearly illustrates these trade-offs. The home composting scenario was estimated to reduce total municipal waste management costs by 15 % compared to a landfill-based reference scenario. This saving was primarily due to the absence of collection, transport, and centralized processing. In contrast, industrial composting, based on weekly door-to-door collection of organic waste, was expected to increase costs by approximately 4 %, mainly due to additional logistical requirements and the operating costs of the central facility [53].

Moreover, from a sociological perspective, decentralized composting shifts waste treatment from a purely technological domain to the educational sphere [61]. By educating its users, taking control of waste treatment on a decentralized scale raises the possibility of both mobilization and environmental awareness. This transforms the users from mere generators into actors in their treatment, demonstrating positive externalities in this domain [62,63]. Maeda et al. [64] observed that Indonesian households in the municipality of Subrabaya adopting this organic recovery route reduced their annual biowaste production by 23 % over a period of three years. In addition, households receiving suitable equipment for waste segregation, on average, sort larger quantities, compared to a situation where they would receive only verbal or written incentives [65,66]. Two key social behaviours can be adopted to illustrate the sociological factors that influence how effectively home composting treats significant amounts of biowaste: (i) the number of
households that use this method of organic recovery, i.e. adherence to this process; and (ii) the assiduity with which practitioners resolve to use it. Loan et al. [67] have demonstrated that knowledge about home composting is the most contributing factor to adherence to this method of waste management. While assiduity among practitioners has a strong positive correlation with already anchored pro-environmental behaviour.

## 3.2. Impact of temperature

Composting is a biological process of decomposition and stabilization of organic matter under certain conditions [58]. The rise in temperature within a compost pile results from the production of heat, resulting from exothermic chemical reactions of two different natures: (i) the oxidation of cellulosic matter, and (ii) the metabolic activity of the microorganisms presents in the environment. The amount of heat produced by biological activity is pre-eminent and serves as an initiator of the cellulosic oxidation process, as it occurs at lower temperatures [59].

In addition to being an intrinsic parameter influencing the decomposition process of organic matter, temperature variations cause changes in microbial communities [68]. Up to a certain limit, the increase in temperature within a compost pile enhances the growth rate and metabolic activity of microorganisms. It also promotes the enzymatic activity of most of the enzymes involved, while increasing the rate of non-enzymatic chemical hydrolysis [69]. Three distinct thermal phases punctuate the process: (i) an initial mesophilic phase (20–45 °C); (ii) a thermophilic phase (45–70 °C); (iii) and a second mesophilic phase of cooling, maturation, and stabilization of the compost obtained [70].

In the biodegradation of some high molecular weight biodegradable polymers (like PLA), the thermophilic step is crucial. This is mainly due to three factors: (i) the increased rate of non-enzymatic chemical hydrolysis reactions; (ii) the influence of thermophilic microorganisms presents during this phase and the impact of high temperatures on enzymatic activity; and (iii) the effect of elevated temperatures on the mobility of polymer molecular chains [71,72]. An elevation of temperature, for example in the case of PLA, close to 60 °C (approximately its glass transition temperature, $T_g$), increases chain flexibility and thereby improves the accessibility of ester bonds to hydrolysis, ultimately

enhancing the biodegradation rate of the polymer [73].

Decentralized composting is often inefficient in reaching and maintaining temperature characteristic of the thermophilic phase for a sufficient period [24,74]. The temperature profile during the decomposition of organic matter is influenced by the surface-to-volume ratio of the waste pile being processed (Fig. 1), the initial chemical composition of the raw material, the aeration rate, profile, humidity, and the possible presence of additives and/or bulking agents [50,75].

The temperature of the disintegrating pile is also influenced by the ambient air temperature [76]. This is particularly true for small volumes of material, where thermal inertia is generally lower, since the mass of compost is not sufficient to effectively retain the heat generated [77,78]. This implies that for temperate climates, the summer season is particularly favourable for the intensification of microbial metabolic activity, in decentralized composting systems [79]. As a result, more intense bio-oxidative and hydrolytic action occurs, and a higher rate of biodegradation of the decomposing material is observed [80].

On the other hand, a marked winter season, as it occurs in the northern part of Europe, and North America, can generate negative temperatures inside a compost pile, thus greatly limiting the speed of chemical reactions, and consequently the progress of the process [81–83].

Thus, seasonal temperature fluctuations significantly influence microbial community dynamics within compost piles. In this context, Hiraishi et al. [84] showed through respiratory quinone profiling that microbial community structure varied seasonally, with a dissimilarity index (D) of up to 0.30. Summer samples contained more partially saturated quinones, suggesting a dominance of thermotolerant bacteria such as *Actinobacteria*. These changes reflect an adaptive microbial succession in response to elevated temperatures, which vary according to the seasons.

Moreover, Shyamala and Belagali [85] further confirmed that microbial activity was significantly higher in summer and autumn, when favourable thermal conditions support a rapid succession of mesophilic, thermotolerant, and thermophilic microorganisms. Conversely, during winter or the rainy season, microbial profiles remained less diverse and organic matter transformation rates slowed due to a more gradual and limited temperature increase.

These seasonal variations in microbial activity and diversity are

<figure><img src="image_2.png" /><figcaption>**Fig. 1.** Schematic illustration of the impact of biowaste volume on temperature in a compost pile. (A): Decentralized compost pile with a higher surface-to-volume ratio than an industrial compost pile generates greater convective heat loss per unit volume. (B): Greater heat loss per unit volume means shorter thermophilic phase duration and lower peak temperature. Created in BioRender. Bou sarkis, A. (2024) BioRender.com/y20z968.</figcaption></figure>
particularly pronounced in small-scale systems such as home composting, compared to centralized facilities. As a result, they can significantly influence the compostability of biodegradable plastics. Indeed, the limited pile volume hinders heat retention and often prevents the establishment of a sustained thermophilic phase, especially in winter [80,86].

This effect largely depends on local climate conditions, particularly the degree of seasonal contrast. A better integration of these regional climatic differences—especially seasonal variability—into biodegradability testing protocols could help bridge the gap often observed between laboratory results and real-world performance of compostable plastics.

It may also be relevant to define specification standards by climatic zones (e.g. Zone A, Zone B), based on regional temperature profiles and seasonal dynamics, to ensure more accurate and context-specific assessments of home-compostable plastics.

Heat generation is relatively independent of the geometric shape of a compost pile, compared to its volume [87]. Therefore, it is not an effective parameter for optimising the composting system for increased heat generation.

On the other hand, the rate and profile of aeration of low-volume composting systems is essential to eliminate excessive moisture present in the pile; because the temperatures characteristic of the thermophilic phase are often not reached due to higher humidity [50,88]. Too little airflow limits the oxidation of the decomposing organic matter, as well as the removal of excess moisture within the compost pile, which has the effect of reducing the biodegradative capacity of the medium [89]. Too much airflow, on the other hand, leads to a loss of energy inside the system by dissipating some of the heat produced by exothermic chemical reactions. An optimal aeration rate should minimize heat loss while maximizing the bio-oxidative activity of microorganisms [90]. Lai et al. [91] determined experimentally that an airflow rate of around 0.4L/min is optimal for a system of 29 × 19.7 × 10.1 (height, length, width; cm), which can contain a maximum of four litters of biowaste. However, these results should be put into perspective because the optimal range of the aeration rate varies due to the nature of the composted biowaste (i.e. particle size and chemical composition) [92]. The same scientists have also demonstrated that bidirectional ventilation allows for faster and more persistent attainment of a higher temperature peak compared to unidirectional air flow. Therefore, the organic matter in the system begins to decompose more quickly. This is because the compost's moisture content and oxygen content are homogenized by a bidirectional airflow coming from both the upper and bottom sections [93].

Thus, achieving and maintaining sufficiently high temperatures within the decomposition environment—specifically within the thermophilic range (45–70 °C)—is a critical factor in enhancing the biodegradability of compostable plastics under home composting conditions. However, sustaining such thermophilic conditions remains challenging in decentralized composting systems, particularly when compost volumes are too low or during colder seasons.

Several practical limitations affect the thermal profile of home composters, such as low thermal inertia, significant heat losses to the environment due to small compost volumes, and excessive moisture accumulation caused by rain exposure when the system is placed outdoors. These constraints highlight the need for adapted strategies to enhance heat generation and retention in such systems [94–96]:

1. Increase the volume of the compost pile: Piles of at least 1 × 1 × 1 m (~1 m³) provide sufficient biomass to accumulate metabolic heat and reduce thermal losses through convective heat transfer.

2. Use insulated or double-walled compost bins, or protect composters from rain and wind using covers, plastic sheeting, or sheltered placement. This reduces cooling and excess moisture, both of which inhibit thermophilic transitions.

3. Alternating "green" materials (rich in nitrogen) with "brown" materials (rich in carbon) helps maintain a C:N ratio between 25:1 and 30:1, which represents an optimal balance for microbial growth and metabolic activity. In addition, the incorporation of coarse bulking agents—such as shredded garden waste, bark, or wood chips—improves air porosity and drainage, thereby helping to regulate moisture and maximize oxygen levels within the compost pile.

4. Ensure sufficient aeration by regularly turning the pile (e.g. once or twice a week) or integrating structural supports such as perforated tubes or raised bins. This stimulates microbial respiration and encourages thermophilic growth by maintaining oxygen levels above 5 %.

5. Avoid excessive watering or compaction: water should be added only as needed to keep the compost moderately moist (~50–60 % moisture content). Excess moisture fills air voids, reduces aeration, and promotes a shift toward anaerobic conditions, which limits heat generation and slows down polymer biodegradation.

6. In temperate or continental climates, it is advisable to adjust the composting schedule according to the seasons: plastic-rich batches should preferably be started during the warmer months, when ambient conditions are more conducive to high levels of microbial metabolic activity.

These adaptations not only improve the efficiency of home compostable plastic biodegradation but also enhance the safety and maturity of the final compost product.

### 3.3. Microorganisms

Microorganisms play a central role in the biodegradation of biodegradable plastics. They actively contribute to the reduction of polymer molecular weight through enzymatic hydrolytic and oxidative activities and subsequently metabolize the resulting low-molecular-weight compounds into transient biomass, water (H₂O), and carbon dioxide (CO₂) under aerobic conditions [97,98]. Without the involvement of the microbial community, the process cannot be considered biodegradation, but rather simple degradation.

Polymer compounds with excessively high molecular weights cannot directly penetrate microbial cell membranes to be metabolized. As a result, extracellular enzymes depolymerize them until their molecular weight is low enough for the resulting oligomers to cross the cell membrane [99]. Thus, the initial cleavage of macromolecular chains represents a key step in the biodegradation of biodegradable polymers.

Both prokaryotic and eukaryotic organisms can biodegrade high molecular weight polymers [100]. More specifically, it is bacteria and fungi that are most actively involved in the decomposition of organic matter within a compost pile [98]. Compared to most terrestrial environments, compost is a particularly rich medium of microorganisms, both in terms of taxonomic diversity and in terms of total abundance. This results in a higher rate of biodegradation of plastic in a compost-like environment compared to a soil-like environment, due to the higher probability and abundance of microbial strain specifically degrading this material [101].

Fungi are referred to as primary decomposers, since they actively participate in the degradation of easily hydrolysable compounds, such as certain sugars and proteins, during the initial mesophilic phase. Fungi capable of metabolizing biodegradable polymers, isolated from compost-like media, include the following genera: *Aspergillus*, *Cladosporium*, *Clonostachys*, *Doratomyces*, *Fusarium*, *Lecanicillium*, *Mortierella*, *Penicillium*, *Thermomyces*, *Verticillium* [101]. While fungi and bacteria initially compete with one another for the substrates necessary for their survival, growth and reproduction, the maximum rates of bacterial-specific growth are eventually exceeding the fungal growth, particularly during the thermophilic phase [61].

Under normal conditions, a biofilm is first formed on the surface of
the plastic material, in several stages [102,103]. This is a critical phase which conditions the effectiveness of the biodegradation process. Bio-film formation on the surface of biodegradable plastics in home-compost environments follows a well-established ecological succession, influenced by both the physicochemical characteristics of the polymer (e.g. roughness, hydrophobicity, surface free energy) and the environmental conditions within the compost matrix (e.g. temperature, humidity) [104]. Among the characteristics of polymers, surface roughness is an important parameter that influences biofilm formation. Upon introduction into compost, the plastic surface is rapidly coated by a conditioning film composed of organic matter, including humic substances, proteins, polysaccharides, and lipids. This film alters the surface energy and promotes microbial adhesion [105,106].

An increase in material surface roughness enhances the effective area available for the initial formation of films composed of organic molecules, such as proteins and polysaccharides, which precede biofilm development [105]. Rougher surfaces also facilitate mechanical interlocking between microbial cells and the polymer substrate, reducing the likelihood of detachment and thereby increasing the microbial residence time [107]. Furthermore, increased roughness can create localized variations in water retention and microenvironmental hydration, which in turn promote more extensive microbial development [107,108]. In summary, compared to a chemically identical but smoother surface, a rougher surface promotes greater microbial adhesion and enrichment.

Within the first day, early colonizers, mainly opportunistic bacteria such as *Pseudomonas*, *Burkholderia*, and *Acidovorax*, as well as filamentous fungi like *Penicillium* and *Aspergillus*, attach reversibly to the conditioned surface [109,110]. As microbial cells begin to produce extracellular polymeric substances (EPS), their adhesion to the surface becomes stronger. This process leads to the formation of microcolonies and is accompanied by the secretion of enzymes capable of initiating the enzymatic cleavage of the polymer's macromolecular chains [110,111].

Over the following days, the biofilm thickens and matures, supporting more diverse and specialized microbial populations [110,111]. As local nutrients become depleted or environmental stress increases, some microbial cells disperse from the biofilm to colonize new substrates, while others persist as active degraders.

Compared to industrial composting, decentralized composting systems are generally more limited in terms of microbial diversity, abundance, and metabolic performance. This limitation is mainly due to the increased variability of environmental conditions in household composters, as well as their small volume, which limits both the duration and intensity of the thermophilic phase [37–39].

The efficiency of microbial-driven biodegradation depends on the availability of active microbial consortia (including bacteria, fungi, and actinomycetes) and on environmental conditions that support their succession and enzymatic potential.

To foster microbial diversity and enzymatic activity under home composting conditions, the following practical measures can be implemented. Some of these recommendations are like those presented at the end of the *Impact of Temperature* section, as the temperature rise within the compost pile strongly depends on the exothermic metabolic activity of the aerobic microorganisms present:

1. Inoculation with effective microorganisms (Bioaugmentation strategy): The addition of microbial consortia that are effective in degrading specific polymers could be a promising strategy to enhance their biodegradability under mesophilic conditions [112–114]. This strategy is discussed in greater detail in the *Bioaugmentation and Biostimulation* section.

2. Promote microbial colonization of biodegradable plastics by mixing them with fresh organic waste. Placing the plastics in direct contact with moist, nitrogen-rich materials (e.g. fruit peels or coffee grounds) helps create a favourable microenvironment and stimulates the growth of microbial populations involved in their biodegradation.

3. Turn the compost pile regularly: frequent aeration (e.g. twice per week) prevents the formation of anaerobic zones and supports aerobic microbial respiration. Moreover, regular turning enhances compost homogeneity, allowing microorganisms to access various substrate surfaces, including those of plastic fragments. This is particularly important because both enzymatic degradation and subsequent bio-assimilation of polymer chains are surface-dependent processes—that is, they require direct contact between the polymer and microbial enzymes or cell membranes [115].

4. Maintain optimal moisture content (~50–60 %): Adequate moisture is essential to support microbial metabolic activity and facilitate enzyme diffusion. If the decomposition environment is too dry, microbial processes slow down or may even stop; if too wet, anaerobic conditions can develop. Sufficient water activity is also crucial to enable enzymatic reactions. Moisture sensors or simple tactile assessments by hand compression can help users monitor and maintain hydration levels compatible with effective biological activity [116].

5. Add structural bulking agents: The incorporation of coarse brown materials (such as shredded leaves, sawdust, or bark) improves porosity and facilitates the diffusion of oxygen and heat throughout the composting pile—both of which are critical for maintaining active aerobic microbial communities [117]. The role of bulking agents is further discussed in the section *Complementary practices to increase the efficiency of the composting process*.

6. Monitor and maintain the pH of the decomposition environment within an appropriate range: most beneficial decomposer microorganisms thrive in a neutral to slightly acidic pH (6.0–7.5). The pH can be adjusted by balancing acidic inputs (e.g. food waste) with alkaline materials (e.g. ashes or crushed eggshells) [116].

7. Avoiding substances that inhibit microbial activity: In excessive amounts, certain organic wastes can exert a bacteriostatic or even bactericidal effect, thereby slowing down the biodegradation process. This is particularly true for citrus fruits, which are rich in polyphenols and can inhibit bacterial activity beyond a certain concentration threshold [117,118].

By implementing these microbial management strategies, users can create more biologically active composting environments, which in turn improve the breakdown of compostable plastics under home conditions.

### 3.4. Norms and standards

The ability of a plastic to biodegrade in a particular environment can be officially recognised in the form of a certification. The latter is granted if the material meets certain requirements, following certain tests being defined by international standard certification methods and are carried out under specific conditions. The process of assessing a material's biodegradability involves applying and comparing it to various biodegradability indicators, the most popular and representative of which is the quantification of carbon dioxide emission over time under aerobic conditions [119].

The use of this form of certification is mandated by national and European regulatory bodies for a product to be designated as compostable, whether in domestic or industrial settings [119]. This requirement stems from legislative efforts to address consumer confusion surrounding bioplastics. Such confusion is largely driven by the visual indistinguishability between compostable and non-compostable items, and by the risk of products being misused for inappropriate end-of-life pathways [120].

These standards fall into two distinct categories. The first are specification standards, which define the requirements that a material must meet to qualify for a specific designation (e.g. NF T51–800). The second are protocol standards, which describe the procedures for performing test methods and specify their domain of validity (e.g. ISO 14855) [47].

There are fewer certifications available for home composting than for its industrial equivalent. Two of them are official standards: AS-5810
(Australia) and NF T51-800 (France). The third is a globally recognized private voluntary certification: OK Compost HOME (TÜV Austria, Belgium). All these certifications require that the tested material must reach a minimum mineralization rate of 90 % under mesophilic conditions (i.e. at temperatures below 30 °C) within 365 days for the test to be considered valid.

Thus, the commonly adopted maximum testing duration of one year in home composting studies primarily reflects regulatory requirements. These time constraints, set by certification schemes, also mirror the typical residence time of compost in domestic settings—ranging from 6 to 12 months—before its application to soil [31].

Soil is a mesophilic environment with a lower biodegradation potential than compost, mainly due to its reduced microbial abundance, diversity, and enzymatic activity [101]. Consequently, if plastic products do not effectively degrade within the timeframe of home composting, they are likely to persist even longer once incorporated into soil, where environmental conditions are less favorable to further degradation [121-123]. This reinforces the importance of ensuring that biodegradation is sufficiently advanced during the composting phase itself, before compost is applied to soil.

ISO 14855-1 and ISO 14855-2 are the primary standardized protocols used to assess the aerobic biodegradation of plastic materials in compost-like environments under mesophilic conditions. To determine the ultimate aerobic biodegradation rate of the tested materials, both standards specify the procedures that must be followed when conducting biodegradation tests under controlled composting conditions. These tests are conducted under laboratory conditions, at a recorded and controlled temperature that is generally maintained constant throughout the experiment. Mature compost from a composting facility

for plant waste or the organic portion of municipal waste serves as a source of nutrients, microbes, and matrix support for the decomposition process. To accurately estimate the amount of CO₂ attributable to the mineralization of the test material—and thus corresponding to its biodegradation rate—the CO₂ emissions from bioreactors containing both the test material and compost are recorded over time. The emissions from control bioreactors containing only compost are then subtracted from this data. To assess the degree of microbial aerobic metabolic activity in the compost, a positive reference material, such as microcrystalline cellulose, is used in accordance with the specified test parameters.

The ecotoxicity of compost obtained from a compostability test should be systematically assessed to ensure its environmental safety, particularly in relation to the potential presence of small non-biodegraded fragments (e.g. nano- and microparticles) and residual degradation by-products (e.g. monomers, dimers, oligomers, and microbial secondary metabolites) [124]. Such assessments typically involve evaluating the response of diverse organisms—including microorganisms, plants, and animals—to direct exposure to the compost or its aqueous extracts.

The NF T51-800 standard requires that compost resulting from an ultimate aerobic biodegradation or disintegration test must not have any adverse effects on human or animal health, nor on the environment. In this context, a phytotoxicity test is mandatory. It involves assessing both the germination rate and the plant biomass of a batch of seedlings, as illustrated in Fig. 2. Recommended species for this type of test include yellow mustard (Sinapis alba), tomato (Solanum lycopersicum), and watercress (Lepidium sativum). These species need to be genetically distant to account for differences in sensitivity. In order to meet the

<figure><img src="image_3.png" /><figcaption>**Fig. 2.** Schematic illustration of a phytotoxicity test for mature compost obtained after an aerobic ultimate biodegradation test, under home composting conditions. Plant seeds are sown in mature compost mixed in variable proportions with a horticultural substrate. They are placed in climate chamber under artificial lighting during their early stages of development. The number of emerged seedlings, fresh and dry biomass are examples of variables measured and compared with a control (compost without plastic biodegradation processes) to assess the compost's phytotoxicity. Additional measurements, such as quantifying the quantity of residual plastic nano microparticles in the compost, can also be carried out. Py-GC-MS: gaseous phase chromatography with pyrolizer module. Created in BioRender. Bou Sarkis, A. (2024) BioRender.com/v90l938.</figcaption></figure>
labelling requirements, these two phytotoxicity indices must reach at least 90 % of the value obtained with compost alone, as specified by the standard.

## 4. Main biodegradable polymers

The main biodegradable polymers available on the market, in terms of overall production capacity by weight, are PLA (40.1 %), starch-based blends (34.7 %), PBAT (8.7 %), PHAs (7.5 %), cellulosic derivatives (6.9 %), and PBS (1.7 %) [125]. Based on their production processes, biodegradable polymers can be categorized into five main groups: (i) polymers obtained through bacterial biosynthesis, using natural or genetically modified strains, such as PHAs; (ii) polymers synthesized from biobased monomers, such as PLA; (iii) polymers directly extracted from biomass, such as starch; (iv) polymers synthesized from petroleum-derived monomers, such as PBS and PBAT; and (v) polymers produced through a combination of the aforementioned approaches [126]. Some polymers fall into multiple categories—for instance, PBS and PBAT, which can currently be synthesized at least partially from biobased monomers at an industrial scale, and are expected to be fully biobased in the coming years [127].

In each environment and for a given polymer, the rate of biodegradation is primarily governed by the material's hydrophilicity and the molecular mobility of its polymer chains, in line with the polymer chain flexibility hypothesis [128–130]. Accordingly, the inherent properties that most strongly influence biodegradability include molecular weight and distribution ($M_w$, $M_n$, PDI), degree of crystallinity ($\chi$ %), glass transition temperature ($T_g$), melting temperature ($T_m$), and overall hydrophilicity. Several strategies, such as the incorporation of fillers or polymer blending, can be employed to increase a polymer's hydrophilicity and/or to reduce its molecular weight, degree of crystallinity, glass transition temperature, and melting temperature. These modifications aim to enhance the plastic's biodegradability under mesophilic composting conditions [73]. Moreover, the incorporation of a natural filler into a biocomposite generally enhances its biodegradability. This is because microorganisms tend to consume the filler first, increasing the material's porosity and thereby accelerating biodegradation by exposing a greater surface area to extracellular hydrolytic enzymes [131]

### 4.1. PLA

Poly (lactic acid) (PLA) is a synthetic aliphatic polyester that is both biobased and biodegradable under industrial composting conditions. It consists of varying ratios of the L- and D-enantiomers of lactic acid and exhibits thermoplastic properties. PLA is currently the most widely produced biodegradable biobased plastic worldwide, with an estimated annual production of 300,000 tons, accounting for 47 % of all biodegradable plastics [132,133]. Its strong potential for future development is largely attributed to its excellent mechanical strength [134]. Moreover, compared to conventional petroleum-based polymers, PLA production is estimated to require 20 % to 55 % less energy [135].

Nevertheless, PLA does not degrade within a sufficiently short time under home composting conditions, even at constant temperatures of 25 °C, 37 °C, or 45 °C [134,121,122,136–138]. Initiation of PLA biodegradation requires temperatures of at least 50 °C, closely related to its glass transition temperature (~60 °C) [123]. Indeed, under mesophilic conditions, PLA remains in a glassy state, resulting in a rigid structure that limits both the accessibility and the reactivity of its ester bonds to hydrolytic enzymes, thereby significantly reducing its biodegradation rate [139,140].

Under aerobic conditions, PLA degrades in two distinct sequential phases. The first is a non-enzymatic hydrolysis-based breakdown triggered by the cleavage of the ester bonds in the polymer chains. The microbial sphere is most probably not involved in this process, which is highly dependent on temperature and humidity and self-catalysed by the accumulation of hydrolysis residues (i.e. $H_3O^+$ or $OH^-$) within the

fragment. Enzymatic hydrolysis of oligomers into lactic acid monomers, which are subsequently bio-assimilated by microbes before being mostly transformed into $CO_2$ and $H_2O$, is the second step. This subsequent phase most likely takes place when the polymer's molecular weight reduces to a range of 10,000–20,000 g/mol [141,142]. The three main types of enzymes responsible for the biodegradation of PLA are lipases, esterases, and alkalases. The microorganisms involved in the biodegradation of PLA are not ubiquitous in the environment [143]. The main microbial genera identified as being important contributors to the biodegradation of PLA are *Pseudomonas*, *Bacillus*, *Paecilomyces*, *Stenotrophomonas*, *Thermomonospora*, and *Thermopolyspora* [144–146].

### 4.2. TPS and TPS-based blends

Starch is a carbohydrate polymer synthesized via photosynthesis in higher plants, where it is stored as semi-crystalline granules. It consists of monosaccharide units linked by glycosidic bonds, classifying it as a polysaccharide. Starch is composed of two homopolymers: linear amylose (~20–25 %), a polymer of D-glucose, and highly branched amylopectin (~75–80 %), a polymer of $\alpha$-D-($1\to4$)-linked glucan chains with $\alpha$-D-($1\to6$) branch points [147,148]. The relative proportions of amylose and amylopectin within the starch polymer depend on the botanical source of the biomass.

Starch does not exhibit flow or softening behaviour because its glass transition temperature is higher than its thermal decomposition temperature. Consequently, it must undergo a plasticization process—also referred to as restructuring—where molecules such as water, sorbitol, or glycerol interact with the hydroxyl groups of starch via hydrogen bonding. Under simultaneous heating and shearing conditions, this interaction induces thermoplastic behaviour, enabling starch to be processed using conventional plastic manufacturing techniques such as extrusion and injection moulding [149].

Starch is considered an attractive biodegradable polymer due to its low cost, low density, and compatibility with other polymers [121]. Moreover, it can biodegrade within a reasonable timeframe across a wide range of mesophilic environments, including home composting systems, marine waters, and soil [150].

However, thermoplastic starch (TPS) exhibits poor mechanical properties (e.g. low tensile strength), as well as limited gas barrier and water vapor resistance, which restrict its applicability in the plastic packaging industry [151,152]. To overcome these inherent limitations, TPS is often blended with other polymers and/or reinforced with natural fibers [153,154]. While TPS is hydrophilic—like most natural polymers—biodegradable polyesters such as PLA, PBAT, PBS, and PCL are hydrophobic. This polarity mismatch frequently results in poor interfacial adhesion, compromising the mechanical integrity and consistency of the resulting materials [155,156].

Due to the low miscibility of TPS with conventional biodegradable polyesters, careful consideration must be given to blend formulation. The main objective of incorporating TPS into polyester-based biodegradable polymers is to reduce production costs while preserving, as much as possible, the mechanical and thermal performance of the material. However, in most cases, the resulting blends exhibit inferior mechanical properties compared to neat polyester, and these properties typically deteriorate further as the TPS content increases [157]

### 4.3. PHA's

Poly(hydroxyalkanoate)s (PHAs) constitutes a family of biobased and biodegradable aliphatic polyesters, comprising over 150 monomer units identified to date. These polymers are biosynthesized by various Gram-positive and Gram-negative bacteria—spanning at least 92 known genera—as intracellular storage compounds, which can be metabolized when required, particularly under abiotic stress conditions [158].

Among PHAs, poly(3-hydroxybutyrate) (PHB) is the most extensively studied and utilized, followed by its copolymer, poly(3-
hydroxybutyrate-co-3-hydroxyvalerate) (PHBV). Due to their biocompatibility and biodegradability, PHAs are widely employed in the biomedical field for applications such as sutures, patches, implants, and tissue engineering scaffolds [159]. They have also gained increasing interest in the single-use plastics sector over the past decade [160].

Despite their potential, the use of PHAs in disposable plastic products remains limited due to their relatively poor mechanical properties—such as low impact resistance, brittleness, and limited flexibility—as well as high production costs and processing challenges [161]. Nonetheless, PHAs are the subject of ongoing research aimed at addressing these technical and economic limitations. Promising strategies include the use of food waste as a low-cost carbon source for microbial fermentation and the development of optimized extrusion techniques to improve processing efficiency [162,163]. Moreover, blending PHAs with other polymers and/or reinforcing them with natural fibers has proven effective in enhancing their mechanical performance, thereby improving their suitability for single-use plastic applications.

Microorganisms capable of degrading PHB are predominantly mesophilic and widely distributed in the environment, accounting for an estimated 0.5 % to 9.6 % of microbial colonies in terrestrial ecosystems [164,165]. As PHB is derived from renewable resources and synthesized by bacterial metabolism, it is readily biodegradable in a variety of mesophilic environments, including marine ecosystems and home composting systems. As such, PHB represents a promising candidate for circular economy approaches [123].

## 4.4. PBAT

Poly (butylene adipate-co-terephthalate) (PBAT) is a semi-crystalline aliphatic-aromatic copolyester synthesized through the polycondensation of 1,4-butanediol with adipic acid and terephthalic acid [166]. The incorporation of aromatic rings from terephthalic acid contributes to PBAT's notable mechanical performance, which is comparable to that of polyethylene (PE), making it a promising biodegradable alternative to conventional polyolefins [167–169].

PBAT contains three different types of ester groups along its carbon backbone: two aliphatic esters derived from adipic acid, and one aromatic ester derived from terephthalic acid. This implies that hydrolysis can theoretically occur at three distinct sites [170].

The aliphatic segments of PBAT confer its biodegradability due to the presence of readily hydrolysable ester bonds. In contrast, the aromatic portions contain ester bonds that are more difficult to hydrolyse due to their proximity to the benzene ring, which limits their reactivity [171, 172]. A terephthalate content ranging from approximately 30 % to 55 % (mol/mol) has been identified as an optimal balance between mechanical performance and biodegradability [131,173].

PBAT undergoes cleavage of its ester bonds in the extracellular environment through enzymatic hydrolysis—primarily mediated by hydrolases such as cutinases, lipases, and esterases—as well as via nonenzymatic hydrolysis. These processes are followed by the microbial assimilation of its monomers, which are metabolized intracellularly through redox reactions feeding into the tricarboxylic acid (TCA) cycle [35]. Initial enzymatic attacks typically target the aliphatic units in the amorphous regions of the polymer, where chain mobility facilitates hydrolysis [131].

PBAT is not compostable under home-composting conditions. Its biodegradation occurs slowly at ambient mesophilic temperatures (25–30 °C) and is mainly driven by mesophilic fungi and bacteria, particularly those belonging to the phyla *Firmicutes* and *Proteobacteria* [174]. Efficient biodegradation requires elevated temperatures. For instance, Wallace et al. [175] demonstrated that in liquid media, degradation is significantly enhanced at 65 °C compared to 50 °C. This effect is likely due to the increased chain mobility at higher temperatures, which approach PBAT's melting point (~120 °C). This enhances the accessibility of ester bonds to hydrolytic enzymes and accelerates the

rate of non-enzymatic chemical hydrolysis.

## 4.5. PBS

Poly (butylene succinate) (PBS) is a semi-crystalline, biodegradable aliphatic polyester with thermoplastic properties. It offers good processability and favourable mechanical performance [176]. PBS is typically synthesized via a two-step polycondensation process involving succinic acid (or its dimethyl ester) and 1,4-butanediol—both of which can be derived from either petroleum-based or renewable sources [177]. PBS exhibits a glass transition temperature of approximately –35 °C and a melting point near 115 °C [178]. Its physical properties are comparable to those of poly (ethylene terephthalate) (PET) [179].

Owing to its excellent elongation at break—often exceeding 900 %— PBS finds applications in a broad range of sectors, including mulch films, compostable bags, catering products, non-woven fabrics, and clothing [180,181]. However, unlike petroleum-derived polymers commonly used for similar purposes (e.g. PE, PAs, PS, PET), PBS remains relatively expensive. This is largely due to the cost-intensive synthesis, which involves electrochemical treatment and the condensation of succinic acid or dimethyl succinate with 1,4-butanediol [182].

PBS is generally not considered to meet the criteria for home compostability [123]. To enhance its biodegradability and mechanical performance, PBS is often copolymerized with other monomers [183]. The polymer's biodegradation behaviour is strongly influenced by its molecular weight and the relative composition of its monomer units. Among PBS-based copolymers, poly (butylene succinate-co-butylene adipate) (PBSA) is the most extensively studied and commercially relevant [184].

As the proportion of adipate units increases, the glass transition and melting temperatures decrease, reaching a minimum at near-equimolar compositions (approximately 62 mol % adipate). Beyond this point, the melting temperature begins to rise again, eventually approaching that of poly (butylene adipate) (PBA) [185]. Compared to PBS, PBSA exhibits enhanced biodegradability across various environments, primarily due to its lower crystallinity—rendering it more susceptible to enzymatic attack—and the greater flexibility of its polymer chains. As a result, PBSA is regarded as a promising candidate to replace conventional polyolefins in numerous applications [186].

# 5. Commercial biodegradable plastics

Marketed plastic products, whether biodegradable or not, are composed of polymer resins and additives [187]. Additives include a wide range of compounds and materials whose purpose is to impart certain desired properties to a plastic product, reduce its cost, or facilitate its manufacturing process [40].

During the processing phase (e.g. thermo-extrusion, injection moulding), constituent can react with each other, resulting in changes in polymer molecular and conformational structure and/or chemical modifications (e.g. nucleating agent, antinucleating agent, modify the molecular weight) [188]. Or alternatively, modify the physical accessibility of the constituents to microorganisms, and/or alter the water diffusivity within the structure [189]. This can influence the biodegradability of the components.

Thus, the biodegradability of marketed plastic products is not necessarily equivalent to the biodegradability of each of their constituents. Consequently, the biodegradability certification process for a plastic product must be carried out for each formulation, even if all its constituents are already considered biodegradable [NF T51–800]. A distinction must be made between the constituents of a plastic (i.e. polymer(s) and additive(s)) and its components (e.g. yoghurt pot with lid). According to French standard [NF T51–800] (2015). In the case of complex plastic products with different components that can be separated by physical methods. All components must meet biodegradability and toxicity requirements for the plastic product to be certified as
domestically compostable. Depending on how easy it is to separate the components of a plastic product, the assessment methods are applied either to the individual components, or to the product itself.

Two distinct cases must be considered when addressing the biodegradability of manufactured biodegradable plastic products under decentralized composting conditions.

(i) If the products are officially certified for home compostability (i. e. NF T51–800, AS 5810, OK Compost HOME), they are expected to biodegrade in accordance with the criteria established by the relevant specification standards under controlled laboratory conditions. Specifically, they must achieve >90 % mineralization at temperatures below 30 °C within 365 days.

(ii) If the products are not certified, their ability to biodegrade under home composting conditions cannot be assured. This applies both to controlled laboratory settings and, even more so, to real-world decentralized composting environments.

In the first case, for biodegradable plastics that comply with a laboratory-based specification standard for home compostability, the key question is whether these products biodegrade in a similar manner under real-world composting conditions. However, the experimental protocols typically used (e.g. ISO 14855) often differ significantly from the conditions found in decentralized systems, raising concerns about the representativeness of the results and their applicability to practical usage scenarios [39].

In most domestic contexts, composting takes place outdoors and is subject to considerable variability: fluctuating temperatures, periods of drying or excessive moisture, limited oxygenation, decomposition media composed of fresh biowaste rather than mature compost, and irregular inputs of organic matter. These factors can significantly slow down the biodegradation of biodegradable plastics [36]. This variability makes direct comparison more uncertain and calls into question the relevance and validity of certification schemes under ordinary home composting conditions.

Several recent studies have indeed evaluated the actual behaviour of certified commercially compostable plastic products under home composting conditions, often reporting outcomes that diverge from the expectations set by certification standards [39,190].

In Australia, Bryson et al. [39] investigated the disintegration of several commercially available dog waste bags certified as home compostable (i.e. AS 5810) in outdoor household composters (25–160 L) filled with organic matter rich in moisture and nitrogen, such as dog feces. Although thermophilic temperatures were reached during the trials (> 60 °C), none of the tested formulations met the disintegration requirements of the standard: i.e. ≥ 90 % mass loss of fragments > 2 mm within 180 days. Some commercial products (labelled HC2 and HC5) even showed an apparent increase in mass (1.51–15.21 %) at the end of the trial, which was attributed to the formation of microbial biofilms on their surface. Macroscopic fragments (> 5 mm) and microplastics (< 2 mm) were found in all decomposition media at the end of the process, raising concerns about the actual biodegradability of these certified products under real-world decentralized composting conditions.

On the other hand, Cafiero et al. [190] compared the performance of compostable bags in two types of household composters: static composters (600 L) and an electromechanical composter (1 m³). In the electromechanical composter, 90 % physical disintegration was achieved within just two months, thanks to more uniform internal temperatures and improved aeration. In contrast, in the static composters, disintegration performance varied greatly depending on the management practices. A degradation rate of 90–96 % was observed after six months, but only when moisture content and compost structure were well controlled by adjusting the ratio of wet material to bulking agents and regularly turning the compost pile. Poor maintenance of the composter significantly slowed down the disintegration of the plastics.

These observations point to a common conclusion: the actual

biodegradability of commercially available plastics certified as home compostable under controlled laboratory conditions varies greatly. This variability depends on the interaction between the product's formulation, local composting conditions (e.g., temperature, moisture, microbial community), and the type of equipment used (e.g., static or electromechanical composter).

In the second case, if a plastic product is labelled as "biodegradable" without having undergone any formal evaluation of its compostability under laboratory conditions, it likely does not meet the criteria required for such a claim. Even more so, it is unlikely to biodegrade under real-world decentralized composting conditions, which are generally less stringent than controlled laboratory environments. In such cases, the "biodegradable" label may be misleading and deceptive, potentially reflecting a form of greenwashing by the company marketing the product.

This is particularly problematic because most consumers do not have a clear understanding of the differences between terms such as bio-plastics, biobased, biodegradable, disintegration, biodegradation, biodegradability, and industrially or domestically compostable [191]. As a result, manufacturers take advantage of these cognitive biases to adopt a fallacious marketing policy to sell their products [192]. False claims by manufacturers lead to a loss of consumer confidence. This can have a negative impact on public policies aimed at combating plastic pollution [193].

In this sense, Moreno et al. [191] analysed the composition of formulations of single-use plastic products sold as "biodegradable" or "ecofriendly" from supermarkets in Brazil (Rio de Janeiro and Sao-Paulo). None of the thirty-six products analysed that were sold as "biodegradable" were effectively biodegradable, as they were made from conventional, non-biodegradable resins (e.g. PE, PS, PP). 93.8 % of them were in fact "oxodegradable" plastics, containing pro-oxidant additives, which we know do not meet the criteria for certification as biodegradable [194].

## 6. Methodology used to select studies testing the home compostability of biodegradable plastics

In total, twenty-seven experiments conducted under controlled home composting conditions and twenty-one under real-life, uncontrolled conditions were included in this review (Tables 1 and 2). These studies were selected based on two main criteria to ensure both relevance and comparability.

The first criterion was that biodegradability had to be assessed either through a respirometric method measuring CO₂ evolution (i.e. mineralization), or via a quantitative approach evaluating material disintegration (e.g. weight loss or reduction in surface area assessed through image analysis). While the former provides direct evidence of microbial degradation, the latter only reflects the physical or chemical breakdown of the material. Studies relying solely on other indirect and/or subjective indicators—such as changes in mechanical properties, thermal stability, molecular weight, infrared spectra, or visual inspection—were excluded, as these methods were deemed insufficiently accurate to reliably assess actual biodegradability. Nevertheless, such data may still provide valuable complementary insights into the degradation process.

The second criterion was that the decomposition medium had to be representative of a composting environment—either mature compost or actively decomposing biowaste—used alone or mixed with other components such as vermiculite or soil. When the medium consisted of mature compost, the experiment was classified as a laboratory-scale study, typically conducted under constant mesophilic temperature conditions. In contrast, when the medium consisted of actively decomposing biowaste under non-constant temperature conditions, the test was classified as a real-life home composting experiment.

Biodegradation rates determined by respirometric measurements, along with degradation levels assessed through material disintegration over the course of composting, were extracted from the selected studies
<caption>**Table 1**</caption>

Compilation of the main results from mineralization and/or disintegration tests on biodegradable plastics, under mesophilic composting conditions at controlled temperature. (l, w, t) (a): length, width, thickness of the samples; χ (%) (b): crystallinity rate of the samples; PHO (c) polyhydroxyoctanoate; kDa (d): kilodalton (1 kDa = 1 000 g.mol⁻¹); Mw (e): molecular mass by weight of the samples; Tg (°C) (f): glass transition temperature of the samples; PI (g): polydispersity index; (d, t) (h): diameter and thickness of the samples; supplier data (i): data provided by the supplier before the implementation phase by the authors of the study; (start/end) (j): the time at which the measurements were taken before or at the end of the composting test; GS2189 (k): commercial blend of PLA-based polymers, and additives; HCa (l): hydrolyzed collagen obtained by alkaline treatment; HCe (m): hydrolyzed collagen obtained by enzymatic treatment; ES (n): egg shell powder; Mater-Bi (o): commercial blend of 60% or more of starch and starch derivatives and of about 40% hydrophilic and biodegradable synthetic resin, made of a lower molecular weight polymer and additives; OMMT (p): hexadecyl trimethylammonium bromide modified natural montmorillonite (modified nano clays); HDI (q): 1,6-hexamethylene diisocyanate, a non-toxic connecting agent; PLLA (r): Poly L-Lactic Acid; aPHA (s): amorphous PHA; TPG (t): thermoplastic gliadin.

<table><thead><tr><th>Decomposition media</th><th>Materials</th><th>Shape &amp; size of samples (l, w, t)<sup>(a)</sup></th><th>T°</th><th>Evaluation methods</th><th>Supplementary</th><th>Results</th><th>Authors</th></tr></thead><tbody><tr><td rowspan="2">Mature compost</td><td>PHBV</td><td rowspan="2">Powder ≤ 500 µm (grain size)</td><td rowspan="2">28°C</td><td rowspan="2">Respirometric test</td><td>χ (%)<sup>(b)</sup>: 85.2</td><td>100 % (130 days)</td><td rowspan="2">Bonnenfant et al. [195]</td></tr><tr><td>PHBV/Galic-Acid</td><td>χ (%): 84.7</td><td>100 % (170 days)</td></tr><tr><td rowspan="10">Mature compost</td><td>PHBV/Quercetin</td><td rowspan="10">Powder ≤ 250µm (grain size)</td><td rowspan="10">28± 2°C</td><td rowspan="10">Respirometric test</td><td>χ (%): 80.4</td><td>100 % (150 days)<br>≈ 100 % (260 days)</td><td rowspan="10">Narancic et al. [123]</td></tr><tr><td>PLA/PCL (80/20; w)</td><td>NA (Failed)</td><td>NA (Failed)</td></tr><tr><td>PLA/PBS (80/20; w)</td><td>NA (Failed)</td><td>NA (Failed)</td></tr><tr><td>PLA/PHB (80/20; w)</td><td>NA (Failed)</td><td>NA (Failed)</td></tr><tr><td>PLA/PHO<sup>(c)</sup> (85/15; w)</td><td>≈ 100 % (365 days)</td><td>≈ 100 % (365 days)</td></tr><tr><td>PHB/PHO (85/15; w)</td><td>≈ 90 % (365 days)</td><td>≈ 100 % (88 days)</td></tr><tr><td>PCL/PHO (85/15; w)</td><td>NA (Failed)</td><td>NA (Failed)</td></tr><tr><td>PCL</td><td>≈ 100 % (90 days)</td><td>≈ 100 % (90 days)</td></tr><tr><td>PHO</td><td>90 % (≈ 100 days)</td><td>Fogasová et al. [196]</td></tr><tr><td>TPS</td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td rowspan="4">Mature compost/ Perlite (1/2; w) + Mineral salts</td><td>PLA/PHB/Citric-Acid-Esters (70/15/15; w)</td><td>Square 50 × 50 × 0.04 mm</td><td>28 °C</td><td>Respirometric test</td><td>≈ 153 kDa<sup>(d)</sup> (M<sub>w</sub><sup>(e)</sup>, start)<br>T<sub>g</sub> (°C)<sup>(f)</sup>: 28</td><td>≈ 198 kDa (M<sub>w</sub>, start)<br>T<sub>g</sub> (°C): 24</td><td></td></tr><tr><td>PLA/PHB/TPS/Citric-Acid-Esters (50/10/25/15; w)</td><td>Square 50 × 50 × 0.035 mm</td><td></td><td></td><td>≈ 219 kDa (M<sub>w</sub>, start)</td><td>≈ 90 % (≈ 82 days)</td><td></td></tr><tr><td>PLA/PHB/Citric-Acid-Esters (65/30/5; w)</td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>PLA</td><td>Square 15 × 15 × 0.2 mm</td><td>25°C</td><td>Respirometric test</td><td>≈ 191 kDa (M<sub>w</sub>, start)<br>= 1.71 T<sub>g</sub> (°C): 61.5</td><td>≈ 90 % (≈ 140 days)</td><td></td></tr><tr><td rowspan="5">Mature compost/ Perlite (1/1; w)</td><td>PLA</td><td>Square 15 × 15 × 0.2 mm</td><td>25°C</td><td>Respirometric test</td><td>≈ 191 kDa (M<sub>w</sub>, start)<br>= 1.71 T<sub>g</sub> (°C): 61.5</td><td>14.9 % (119 days)<br>19.5 % (119 days)</td><td>Sedničková et al. [134]</td></tr><tr><td>PLA</td><td>Circular 10 × 3 mm (d, t)<sup>(h)</sup></td><td>25°C</td><td>Weight loss</td><td>30 kDa (M<sub>w</sub>, supplier data<sup>(i)</sup>)</td><td>0 % (10 months)</td><td>Al Hosni et al. [121]</td></tr><tr><td>PCL</td><td></td><td></td><td></td><td>80 kDa (M<sub>w</sub>, supplier data)</td><td>≈ 80 % (10 months)</td><td></td></tr><tr><td>PHB</td><td></td><td></td><td></td><td>55 kDa (M<sub>w</sub>, supplier data)</td><td>≈ 50–55 % (10 months)</td><td></td></tr><tr><td>PBS</td><td></td><td></td><td></td><td>40 kDa (M<sub>w</sub>, supplier data)</td><td>≈ 0 % (10 months)</td><td></td></tr><tr><td rowspan="5">Soil/Compost (90/10; w)</td><td>PLA</td><td>Square 15 × 15 × 0.2 mm</td><td>25°C</td><td>Respirometric test</td><td>≈ 0 % (10 months)</td><td>0 % (10 months)</td><td></td></tr><tr><td>PCL</td><td></td><td></td><td></td><td>≈ 80 % (10 months)</td><td>≈ 75 % (10 months)</td><td></td></tr><tr><td>PHB</td><td></td><td></td><td></td><td>≈ 5–10 % (10 months)</td><td>≈ 5–10 % (10 months)</td><td></td></tr><tr><td>PBS</td><td></td><td></td><td></td><td>50 % (6.1 weeks)</td><td></td><td></td></tr><tr><td>TPS</td><td>Rectangular 10 × 22 × 0.6± 0.1 mm</td><td>25°C</td><td>Respirometric test</td><td>590/548 kDa (M<sub>w</sub>) PI = 2.8/2.5 (start/end)<sup>(j)</sup> χ (%) = 58</td><td>50 % (13.5 weeks)</td><td>Arcos-Hernandez et al. [197]</td></tr><tr><td></td><td>PHBV (HV: 12 %)</td><td></td><td></td><td></td><td>590/495 kDa (M<sub>w</sub>) PI = 2.8/2.8 (start/end) χ (%): 39</td><td>50 % (16.4 weeks)</td><td></td></tr><tr><td></td><td>PHBV (HV: 43 %)</td><td></td><td></td><td></td><td>520/426 kDa (M<sub>w</sub>) PI = 3.3/2.7 (start/end) χ (%): 14</td><td>50 % (17.6 weeks)</td><td></td></tr><tr><td></td><td>PHBV (HV: 47 %)</td><td></td><td></td><td></td><td>590/466 kDa (M<sub>w</sub>) PI = 2.9/2.5 (start/end) χ (%): 29</td><td>50 % (14.5 weeks)</td><td></td></tr><tr><td></td><td>PHBV (HV: 52 %)</td><td></td><td></td><td></td><td>630/510 kDa (M<sub>w</sub>) PI = 2.7/2.4 (start/end) χ (%): 42</td><td>50 % (17.3 weeks)</td><td></td></tr><tr><td></td><td>PHBV (HV: 64 %)</td><td></td><td></td><td></td><td>630/497 kDa (M<sub>w</sub>) PI = 2.7/2.5 (start/end) χ (%): 40</td><td>50 % (17.3 weeks)</td><td></td></tr></tbody></table>

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<table><caption>**Table 1 (continued)**</caption><thead><tr><th>Decomposition media</th><th>Materials</th><th>Shape & size of samples (l, w, t)<sup>(a)</sup></th><th>T°</th><th>Evaluation methods</th><th>Supplementary</th><th>Results</th><th>Authors</th></tr></thead><tbody><tr><td rowspan="2">Composted soil</td><td>Nylon4</td><td>Rectangular 10 × 50 × 0.05–0.07 mm</td><td>25°C</td><td>Weight loss</td><td>140 kDa (M<sub>w</sub>, start)</td><td>100 % (≈ 45 days)</td><td>Hashimoto et al. [198]</td></tr><tr><td>Nylon6<br/>Nylon4/Nylon6 (30/70; w)</td><td></td><td></td><td></td><td>37 kDa (M<sub>w</sub>, start)</td><td>0 % (450 days)<br/>≈ 30 % (≈ 180 days)</td><td></td></tr><tr><td rowspan="5">Mature compost</td><td>PLA</td><td>Rectangular 80 × 10 × 2–4 mm</td><td>25°C</td><td>Weight loss</td><td></td><td>3.08 % (120 days)</td><td>Alimuzzaman et al. [136]</td></tr><tr><td>Flax fibres</td><td></td><td></td><td></td><td></td><td>91.41 % (120 days)</td><td></td></tr><tr><td>PLA/Flax fibres (60/40; w)</td><td></td><td></td><td></td><td></td><td>6.03 % (120 days)</td><td></td></tr><tr><td>PLA/Flax fibres (50/50; w)</td><td></td><td></td><td></td><td></td><td>≈ 12 % (120 days)</td><td></td></tr><tr><td>PLA/Flax fibres (40/60; w)</td><td></td><td></td><td></td><td></td><td>19.77 % (120 days)</td><td></td></tr><tr><td rowspan="15">Composted soil</td><td>PBS</td><td>Rectangular 30 × 40 × NA mm</td><td>30± 2°C</td><td>Weight loss</td><td></td><td>31.4 % (180 days)</td><td>Liu et al. [199]</td></tr><tr><td>Jute fibres</td><td></td><td></td><td></td><td></td><td>24.7 % (180 days)</td><td></td></tr><tr><td>PBS/Jute fibres (90/10; w)</td><td></td><td></td><td></td><td></td><td>62.5 % (180 days)</td><td></td></tr><tr><td>PBS/Jute fibres (80/20; w)</td><td></td><td></td><td></td><td></td><td>54.8 % (180 days)</td><td></td></tr><tr><td>PBS/Jute fibres (70/30; w)</td><td></td><td></td><td></td><td></td><td>47.3 % (180 days)</td><td></td></tr><tr><td>PBS/Jute fibres (90/10; w) (untreated)</td><td></td><td></td><td></td><td></td><td>≈ 60–65 % (180 days)</td><td></td></tr><tr><td>PBS/Jute fibre (90/10; w) (alkaline treatment)</td><td></td><td></td><td></td><td></td><td>≈ 50–55 % (180 days)</td><td></td></tr><tr><td>PBS/Jute fibres (90/10; w) (coupling agent treatment)</td><td></td><td></td><td></td><td></td><td>≈ 40 % (180 days)</td><td></td></tr><tr><td>PBS/Jute fibres (80/20; w) (woven fibres)</td><td></td><td></td><td></td><td></td><td>≈ 65 % (180 days)</td><td></td></tr><tr><td>PBS/Jute fibres (80/20; w) (non-woven fibres)</td><td></td><td></td><td></td><td></td><td>≈ 60 % (180 days)</td><td></td></tr><tr><td>PBS/Jute fibres (80/20; w) (bulk fibres)</td><td></td><td></td><td></td><td></td><td>≈ 40 % (180 days)</td><td></td></tr><tr><td>Poultry manure/ Wheat straw/ Tomato crop residues (1/0.2/2; w)</td><td>GS2189<sup>(k)</sup> /Biochar (95/5; w)</td><td>Rod support (Tomato crop)</td><td>30°C</td><td>Weight loss</td><td></td><td>5.48 % (48 weeks)</td><td>Malińska et al. [200]</td></tr><tr><td rowspan="3">Mature compost (from olive oil mill waste)</td><td>PBSA</td><td>Square 10 × 10 × 0.2 mm</td><td>25°C</td><td>Respirometric test</td><td></td><td>71.7 % (55 days)</td><td>Altieri et al. [187]</td></tr><tr><td>PBSA/HCa<sup>(l)</sup> (95/5; w)</td><td></td><td></td><td></td><td></td><td>70.9 % (55 days)</td><td></td></tr><tr><td>PBSA/HCa (80/20; w)</td><td></td><td></td><td></td><td></td><td>56.6 % (55 days)</td><td></td></tr><tr><td rowspan="2">Composted soil</td><td>PBSA/HCe<sup>(m)</sup> (80/20; w)</td><td></td><td></td><td></td><td></td><td>48.4 % (55 days)</td><td></td></tr><tr><td>PCL</td><td>Square 10 × 10 × 3 mm</td><td>43°C</td><td>Weight loss</td><td>56 kDa (M<sub>w</sub>, supplier data)</td><td>≈ 40 % (8 weeks)</td><td>González Petit et al. [201]</td></tr><tr><td rowspan="2">Mature compost</td><td>PCL/ES (n) (1/1; w)</td><td></td><td></td><td></td><td></td><td>≈ 40 % (8 weeks)</td><td></td></tr><tr><td>Polyurethane</td><td>Square 40 × 40 × 1 mm</td><td>25°C</td><td>Weight loss</td><td></td><td>≈ 40 % (12 weeks)</td><td>Zafar et al. [202]</td></tr><tr><td rowspan="6">Compost/Fresh soil (1/1; w)</td><td>PBS/TPS (1/1; w)</td><td>Rectangular 100 × 150 × 0.05 mm</td><td>25°C</td><td>Weight loss</td><td>TPS: 123 kDa PBS: 161 kDa (M<sub>w</sub>, start)</td><td>79.2 % (30 days)</td><td>Fahrngruber et al. [203]</td></tr><tr><td>PBS/TPS (1/1; w) + Compatibilizer (0.5 %): Starch</td><td>Rectangular 100 × 150 × 0.05 mm</td><td></td><td></td><td>TPS: 119 kDa PBS: 159 kDa (M<sub>w</sub>, start)</td><td>76.6 % (30 days)</td><td></td></tr><tr><td>PBS/TPS (1/1; w) + Compatibilizer (1 %): Starch</td><td>Rectangular 100 × 150 × 0.045 mm</td><td></td><td></td><td>TPS: 96 kDa PBS: 178 kDa (M<sub>w</sub>, start)</td><td>65.5 % (30 days)</td><td></td></tr><tr><td>PBS/TPS (1/1; w) + Compatibilizer (0.5 %): TPS</td><td>Rectangular 100 × 150 × 0.045 mm</td><td></td><td></td><td>TPS: 116 kDa PBS: 171 kDa (M<sub>w</sub>, start)</td><td>69.2 % (30 days)</td><td></td></tr><tr><td>PBS/TPS (1/1; w) + Compatibilizer (1 %): TPS</td><td>Rectangular 100 × 150 × 0.04 mm</td><td></td><td></td><td>TPS: 69 kDa PBS: 177 kDa (M<sub>w</sub>, start)</td><td>77.7 % (30 days)</td><td></td></tr><tr><td>Mature compost</td><td>Mater-Bi<sup>(o)</sup></td><td>Square 150 × 150 × NA mm</td><td>25°C</td><td>Weight loss</td><td></td><td>43 % (90 days)</td><td>Accinelli et al. [204]</td></tr></tbody></table>

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<table><caption>Table 1 (continued)</caption><thead><tr><th>Decomposition media</th><th>Materials</th><th>Shape & size of samples (l, w, t)<sup>(a)</sup></th><th>T°</th><th>Evaluation methods</th><th>Supplementary</th><th>Results</th><th>Authors</th></tr></thead><tbody><tr><td rowspan="2">Mature compost</td><td>TPS</td><td>Square 20 × 20 × 0.245 mm</td><td rowspan="2">28°C</td><td rowspan="2">Respirometric test</td><td>χ (%): 3.58 Moisture content (%): 21.95</td><td>93 % (75 days)</td><td rowspan="2">Díaz-Díaz et al. [205]</td></tr><tr><td>TPS/Salicylic-Acid</td><td></td><td>χ (%): 6.6 Moisture content (%): 20.6</td><td>85 % (75 days)</td></tr><tr><td rowspan="2">Vermicompost</td><td>PLA</td><td>Rectangular 100 × 20 × 0.03 mm</td><td rowspan="2">27°C</td><td rowspan="2">Respirometric test</td><td>PLA: 211 kDa PHB: 426 kDa (M<sub>w</sub>, supplier data)</td><td>≈ 90 % (≈ 60 days)</td><td rowspan="2">Jandas et al. [206]</td></tr><tr><td>PLA/PHB/Maleic-Anhydride (70/30/7; w)</td><td></td><td>≈ 90 % (≈ 53 days)</td></tr><tr><td rowspan="7">Garden soil (high peat content)</td><td>PLA</td><td>Rectangular 1.6–1.8 mm (thickness)</td><td rowspan="7">30°C</td><td rowspan="7">Weight loss</td><td>111 kDa (M<sub>w</sub>, supplier data)</td><td>2.84 % (6 mois)</td><td rowspan="7">Latos-Brozioand Masek [137]</td></tr><tr><td>PLA/Beta-carotene (1/1; w)</td><td></td><td>3.22 % (6 mois)</td></tr><tr><td>PLA/Juglone (1/1; w)</td><td></td><td>0.6 % (6 mois)</td></tr><tr><td>PLA/Morin (1/1; w)</td><td></td><td>≈ 1.3 % (6 mois)</td></tr><tr><td>PLA/Curcumin (1/1; w)</td><td></td><td>≈ 3.3 % (6 mois)</td></tr><tr><td>P(3,4 HB)</td><td></td><td>520 kDa (M<sub>w</sub>, supplier data) 12 mol % 4-hydroxybutyrate</td><td>5.64 % (6 mois)</td></tr><tr><td>P(3,4 HB)/Beta-carotene (1/1; w)</td><td></td><td>≈ 0.8 % (6 mois)</td></tr><tr><td rowspan="5">Manure compost/Soil/ Sand (1/1/1; w)</td><td>PLA/Morin (1/1; w)</td><td></td><td></td><td></td><td>3.67 % (6 mois)</td><td></td><td></td></tr><tr><td>P(3,4 HB)/Juglone (1/1; w)</td><td></td><td></td><td></td><td>0.4 % (6 mois)</td><td></td><td></td></tr><tr><td>P(3,4 HB)/Curcumin (1/1; w)</td><td></td><td></td><td></td><td>0.8 % (6 mois)</td><td></td><td></td></tr><tr><td>PHB</td><td>Square 20 × 20 × 0.1–0.12 mm</td><td rowspan="5">25–30°C</td><td rowspan="5">Weight loss</td><td>133.4 kDa (M<sub>w</sub>, PI = 2.3 (supplier data)</td><td>≈ 33 % (21 days)</td><td rowspan="5">Saad et al. [207]</td></tr><tr><td>PHB-UV3 (3 h UV exposure)</td><td></td><td>≈ 41 % (21 days)</td></tr><tr><td>PHB-UV6 (6 h UV exposure)</td><td></td><td>≈ 45 % (21 days)</td></tr><tr><td>PHB-UV9 (9 h UV exposure)</td><td></td><td>≈ 54 % (21 days)</td></tr><tr><td>UPCLI-HB (PCL/ PHB/HDI<sup>(q)</sup>)(51.5/ 39.7/8.8; w)</td><td></td><td>56.9 kDa (M<sub>w</sub>, start), PI = 2.2</td><td>≈ 23 % (21 days)</td></tr><tr><td rowspan="5">Raw waste*</td><td>UPCLI-HB (PCL/ PHB/HDI)(51.5/ 39.7/8.8; w)</td><td></td><td></td><td></td><td>≈ 38 % (21 days)</td><td></td><td></td></tr><tr><td>UPCLI-HB (PCL/ PHB/HDI)(51.5/ 39.7/8.8; w)</td><td></td><td></td><td></td><td>≈ 44 % (21 days)</td><td></td><td></td></tr><tr><td>UPCLI-HB (PCL/ PHB/HDI)(51.5/ 39.7/8.8; w)</td><td></td><td></td><td></td><td>≈ 48 % (21 days)</td><td></td><td></td></tr><tr><td>PLA</td><td>Rectangular 10 × 20 × 1 mm</td><td rowspan="5">37°C</td><td rowspan="5">Weight loss</td><td>χ (%): 1.9</td><td>≈ 1 % (24 weeks)</td><td rowspan="5">Sourkouni et al. [138]</td></tr><tr><td>PLA-US20kHz-6 (ultrasonic waves: 20kHz/6 h)</td><td></td><td>≈ 0.3 % (24 weeks)</td></tr><tr><td>PLA-US860kHz-6</td><td></td><td>≈ 1.6 % (24 weeks)</td></tr><tr><td>PLA-UV6 (exposure to UV for 6 h)</td><td></td><td>≈ 0.9 % (24 weeks)</td></tr><tr><td>PLA-US20kHz-UV6</td><td></td><td>≈ 1.1 % (24 weeks)</td></tr><tr><td rowspan="3">Vermicompost (horse manure/Household waste)</td><td>Alginate/Glycerol (95/5; w)</td><td>Rectangular 40 × 50 × NA mm</td><td rowspan="3">30°C</td><td rowspan="3">Weight loss</td><td>Alginate/Glycerol (95/5; w) + EP (2%; w)</td><td>≈ 0.6 % (24 weeks)</td><td rowspan="3">Villanueva et al. [208]</td></tr><tr><td>Alginate/Glycerol (95/5; w) + EP (4%; w)</td><td></td><td>≈ 100 % (21 days)</td></tr><tr><td></td><td></td><td>≈ 100 % (21 days)</td></tr></tbody></table>

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<table><caption>**Table 1 (continued)**</caption><thead><tr><th>Decomposition media</th><th>Materials</th><th>Shape & size of samples (l, w, t)<sup>(a)</sup></th><th>T°</th><th>Evaluation methods</th><th>Supplementary</th><th>Results</th><th>Authors</th></tr></thead><tbody><tr><td rowspan="5">Mature compost</td><td>Alginate/Glycerol (95/5; w) + EP (6%; w)</td><td rowspan="5">Rectangular 10 × 30 × 2 mm</td><td rowspan="5">45°C</td><td rowspan="5">Weight loss</td><td rowspan="5">162/128 kDa (M<sub>w</sub>) PI = 1.7/2.1 (start/end)</td><td>≈ 100 % (14 days)</td><td rowspan="5">Wilfred et al. [122]</td></tr><tr><td>PLA</td><td>0.7 % (14 days)</td></tr><tr><td>PLA/Starch (90/10; w)</td><td>3.7 % (14 days)</td></tr><tr><td>PLA/Starch (75/25; w)</td><td>4.3 % (14 days)</td></tr><tr><td>PLA/Starch (50/50; w)</td><td>14.9 % (14 days)</td></tr><tr><td>PLA/Starch (25/75; w)</td><td>24.5 % (14 days)</td></tr><tr><td>PLA/Starch (10/90; w)</td><td>44.8 % (14 days)</td></tr><tr><td>Sterilised compost (non-inoculated)</td><td>PLA</td><td>Powder ≈ 0.05 mm (grain size)</td><td rowspan="2">37°C</td><td rowspan="2">Respirometric test</td><td rowspan="2">188.7 kDa (M<sub>w</sub>, start) PI = 3.18</td><td>≈ 2 % (40 days)</td><td rowspan="2">Jeon and Kim [145]</td></tr><tr><td>Sterilised compost (with PLA-degrading strains)</td><td>PLA</td><td>Rectangular 0.3 mm (thickness)</td><td>≈ 31 % (40 days)</td></tr><tr><td rowspan="5">Mature compost/ Sand/Rabbit food</td><td>PLA-UV8 (exposure to UV for 8 h)</td><td rowspan="5">Powder ≤ 0.3 mm</td><td rowspan="5">30°C</td><td rowspan="5">Respirometric test</td><td>97.8 kDa (M<sub>w</sub>, start) PI = 2.15</td><td>≈ 45 % (40 days)</td><td rowspan="5">Han et al. [209]</td></tr><tr><td>PLA-UV16 (exposure to UV for 16 h)</td><td>66.3 kDa (M<sub>w</sub>, start) PI = 2.25</td><td>≈ 38 % (40 days)</td></tr><tr><td>PLLA<sup>(r)</sup></td><td>PLLA: 233 kDa, PI = 1.99 aPHA<sup>(s)</sup>: 603 kDa, PI = 2.79 (M<sub>w</sub>, supplier data) χ (%) : 1.9</td><td>≈ 1 % (45 days)</td></tr><tr><td>Quenched PLLA</td><td>χ (%): ≈ 0.1</td><td>≈ 22 % (45 days)</td></tr><tr><td>PLLA/aPHA (90/10; w)</td><td>71.9 % (45 days)</td></tr><tr><td rowspan="5">Mature compost/ Vermiculite (4/1; w)</td><td>PLLA/aPHA (70/30; w)</td><td rowspan="2">≥ 0.3 mm</td><td rowspan="2"></td><td rowspan="2"></td><td rowspan="2"></td><td>≈ 22 % (45 days)</td><td rowspan="5">Mayekar and Auras [210]</td></tr><tr><td>PLLA/aPHA (50/50; w)</td><td>82.7 % (45 days)</td></tr><tr><td>PLLA/aPHA (90/10; w) (annealing)</td><td rowspan="2">≤ 0.3 mm</td><td rowspan="2"></td><td rowspan="2"></td><td rowspan="2"></td><td>4.2 % (45 days)</td></tr><tr><td>PLLA/aPHA (70/30; w)(annealing)</td><td>31.6 % (45 days)</td></tr><tr><td>PLLA/aPHA (50/50; w) (annealing)</td><td>≥ 0.3 mm</td><td></td><td></td><td></td><td>56.2 % (45 days)</td></tr><tr><td rowspan="3">Mature compost/ Vermiculite (4/1; w)</td><td>PLA</td><td>Pellets</td><td rowspan="3">37°C</td><td rowspan="3">Respirometric test</td><td>χ (%): 28.3/31.4 (start/ end)</td><td>Negative rate (≈ -30 % in 180 days)</td><td rowspan="3">Aragón-Gutiérrez et al. [211]</td></tr><tr><td>PLA + Fe (additiv in compost; 8 g)</td><td>χ (%): 28.3/40.9 (start/ end)</td><td>Negative rate (≈ -15 % in 180 days)</td></tr><tr><td>PLA + Gelatin/Fe (additivs in compost; 8 g)</td><td>χ (%): 57.1, Water absorption(%): 0.2</td><td>Negative rate (≈ -150 % in 180 days)</td></tr><tr><td rowspan="3">Mature compost/ Vermiculite (4/1; w)</td><td>PCL</td><td>Powder</td><td rowspan="3">25 ± 5°C</td><td rowspan="3">Respirometric test, Weight loss</td><td>χ (%): 57.1, Water absorption(%): 0.2</td><td>≈ 33 % (210 days)</td><td rowspan="3">Aragón-Gutiérrez et al. [211]</td></tr><tr><td>PCL/TPG<sup>(t)</sup> (75/25; w)</td><td>χ (%): 65.6 Water absorption (%): 8.1</td><td>79 %/≈ 90 % (210/ 120 days)</td></tr><tr><td>PCL/TPG (50/50; w)</td><td>χ (%): 83.1 Water absorption (%): 16.3</td><td>86 %/≈ 90 % (210/ 120 days)</td></tr><tr><td rowspan="4">Mature compost/ Vermiculite (4/1; w)</td><td>PLA</td><td>Pellets</td><td rowspan="4">37°C</td><td rowspan="4">Respirometric test</td><td>χ (%): 28.3/31.4 (start/ end)</td><td>Negative rate (≈ 30 % in 180 days)</td><td rowspan="4">Mayekar and Auras [212]</td></tr><tr><td>PLA + Gelatin (additiv in compost; 8 g)</td><td>χ (%): 28.3/41.9 (start/ end)</td><td>56.4 % (180 days)</td></tr><tr><td>PLA + Ethyl lactate (additiv in compost; 8 g)</td><td>χ (%): 28.3/38.4 (start/ end)</td><td>105 % (180 days)</td></tr><tr><td>PLA + Skim milk (additiv in compost; 8 g)</td><td>χ (%): 28.3/39.5 (start/ end)</td><td>87.7 % (180 days)</td></tr></tbody></table>
<caption>**Table 2**</caption>

Compilation of the main results from disintegration tests on biodegradable plastics, under home-composting, at non-controlled temperature. (l, w, t) (a): (length, width, height); Mater-Bi (b): commercial blend of starch, ethylene vinyl alcohol and additives; Bioflex-219F (c): commercial blend of PLA, co-polyester and additives; Mater-Bi (d): commercial blend of 60 % or more of starch and starch derivatives and of about 40 % hydrophilic and biodegradable synthetic resin, made of a lower molecular weight polymer and additives; MA-g-SEBS (e): Maleic anhydride grafted interfacial modifier; DS: (f): degree of substitution; CS (g): chitosan; CRE: (h): Chinese chive root extract; ES (i): egg shell powder; GS2189 (j): commercial blend of PLA, starch and additives.

<table><thead><tr><th>Decomposition media</th><th>Type of composter</th><th>Materials</th><th>Shape & size of samples (l, w, t)<sup>(a)</sup></th><th>T° (min/max)</th><th>Results (Weight loss)</th><th>Authors</th></tr></thead><tbody><tr><td>Kitchen waste/Bulking agent (10/1; w)</td><td rowspan="3">Compost bin (1200 L)(lateral and basal vent)</td><td>Mater-Bi<sup>(b)</sup></td><td rowspan="3">Plastic bag 50 × 37 cm × 50 µm</td><td>≈ 18–55°C (cumulative T° +++)</td><td>77 % (45 days)<br>96 % (180 days)</td><td rowspan="3">Cafiero et al. [191]</td></tr><tr><td>Plant waste/Bulking agent (10/1; w) + Fresh grass</td><td></td><td>≈ 18–53°C (cumulative T° +++)</td><td>85 % (45 days)<br>90 % (180 days)</td></tr><tr><td>Kitchen waste</td><td></td><td>≈ 18–47°C (cumulative T° +)</td><td>52 % (45 days)<br>95 % (180 days)</td></tr><tr><td>Rabbit manure/Grass clippings/ Household biowaste</td><td>Wooden compost bin (800 & 2000 L)</td><td>TPS<br>TPS Bioflex-219F<sup>(c)</sup></td><td>Square 10 × 10 × NC mm</td><td>13.6–71.4 °C</td><td>9.4 % (14 weeks)<br>0 % (14 weeks)<br>1.7 % (14 weeks)</td><td>Vaverková et al. [213]</td></tr><tr><td>Kitchen waste/Garden waste/Paper/ Water(38/45/0.04/0.13; w)</td><td>Drum composter (200 L)</td><td>Mater-Bi<sup>(d)</sup> (60/40; w)</td><td>Rectangular 2.5 × 5 × NC mm</td><td>22–63 °C</td><td>26.9 % (72 days)<br>90 % (90 days)</td><td>Mohee et al. [214]<br>Akonda et al. [215]</td></tr><tr><td>Mature compost</td><td>Flower pot</td><td>Flax fibres</td><td>NA</td><td></td><td>10 % (120 days)<br>27 % (120 days)</td><td></td></tr><tr><td rowspan="5">NA</td><td rowspan="5">Compost bin</td><td>PLA</td><td>NA</td><td></td><td>19 % (120 days)</td><td></td></tr><tr><td>PLA/Nonwoven-Mats (Flax)</td><td>Rectangular 60 × 25 × 2.1 mm</td><td></td><td></td><td></td></tr><tr><td>PLA/Unidirectional-Fiber (Flax)</td><td>Rectangular 60 × 25 × 2.3 mm</td><td></td><td></td><td></td></tr><tr><td>PHB (A)</td><td>Square 85 × 85 × 2 mm</td><td>5–30 °C</td><td>4 % (150 days)<br>Composter (A)</td><td rowspan="5">Mergaert et al. [216]</td></tr><tr><td>PHBV (10 % HV) (A)</td><td></td><td>5–30 °C</td><td>6 % (150 days)<br>Composter (A)</td></tr><tr><td>PHBV (20 % HV) (A)</td><td></td><td>5–30 °C</td><td>70 % (150 days)<br>Composter (A)</td></tr><tr><td>PHBV (10 % HV) (B)</td><td></td><td>12.5–15 °C</td><td>5 % (150 days)<br>Composter (B)</td></tr><tr><td>PHBV (10 % HV) (C)</td><td></td><td>12.5–30 °C</td><td>14 % (150 days)<br>Composter (C)</td></tr><tr><td rowspan="5">Organic waste</td><td rowspan="5">Drum composter with vents</td><td>PBAT/TPS (43/57; w) + Glycerol</td><td>Square 50 × 50 × 0.18 mm</td><td>14.9–57 °C</td><td>≈ 16 % (28 days)</td><td rowspan="5">Taiatele et al. [217]</td></tr><tr><td>PLA/TPS (50/50; w) + Glycerol</td><td>Square 50 × 50 × 0.17 mm</td><td></td><td>≈ 25 % (28 days)</td></tr><tr><td>PVA/TPS (25/75; w) + Glycerol</td><td>Square 50 × 50 × 0.21 mm</td><td></td><td>≈ 54 % (28 days)</td></tr><tr><td>PHB</td><td>Dog bone</td><td>40–62 °C</td><td>50 % (12 weeks)<br>≈ 58 % (12 weeks)<br>≈ 60 % (12 weeks)<br>68.11 % (12 weeks)<br>≈ 66 % (12 weeks)<br>≈ 69 % (12 weeks)<br>84.11 % (12 weeks)<br>≈ 51 % (12 weeks)<br>≈ 58 % (12 weeks)<br>68.15 % (12 weeks)</td></tr><tr><td>PHB/Hemp (90/10; w)</td><td></td><td></td><td></td></tr><tr><td rowspan="10">Mature compost/Household biowaste/ Green waste</td><td rowspan="10">Wooden composter (3 000 L)</td><td>PHB/Hemp (80/20; w)</td><td></td><td></td><td></td><td></td></tr><tr><td>PHB/Hemp (70/30; w)</td><td></td><td></td><td></td><td></td></tr><tr><td>PHB/Jute (90/10; w)</td><td></td><td></td><td></td><td></td></tr><tr><td>PHB/Jute (80/20; w)</td><td></td><td></td><td></td><td></td></tr><tr><td>PHB/Jute (70/30; w)</td><td></td><td></td><td></td><td></td></tr><tr><td>PHB/Lyocel (90/10; w)</td><td></td><td></td><td></td><td></td></tr><tr><td>PHB/Lyocel (80/20; w)</td><td></td><td></td><td></td><td></td></tr><tr><td>PHB/Lyocel (70/30; w)</td><td></td><td></td><td></td><td></td></tr><tr><td>Cassava starch/Glycerol/ Carboxymethyl-cellulose + Turmeric oil (0,5 µl)</td><td>Square 20 × 20 × 1.8 mm</td><td>≈ 40°C (Average T°)</td><td>≈ 1 % (70 days)</td><td>Fazita et al. [219]</td></tr><tr><td>Cassava starch/Glycerol/ Carboxymethyl-cellulose + Turmeric oil (30 µl)</td><td>Rectangular 80 × 25 × 2 mm</td><td>22.7–38.7°C</td><td>34.23 % (2 months)</td><td>Mustapha et al. [220]</td></tr></tbody></table>

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<table><caption>**Table 2 (continued)**</caption><thead><tr><th>Decomposition media</th><th>Type of composter</th><th>Materials</th><th>Shape & size of samples (l, w, t)<sup>(a)</sup></th><th>T° (min/max)</th><th>Results (Weight loss)</th><th>Authors</th></tr></thead><tbody><tr><td rowspan="3">Composted soil</td><td rowspan="3">Soil/Wood sawdust/Urban waste (≈ 44 % w) + Pineapple juice<br>Household biowaste/Treatment plant sludge (2.5/1; w)</td><td>Cassava starch/Glycerol/ Carboxymethyl-cellulose + Turmeric oil (0,5 µl)</td><td>Rectangular 80 × 25 × 0.5 mm</td><td rowspan="3">20 ± 2°C<br>(Average T°)</td><td>53.01 % (2 months)</td><td rowspan="3">Merino et al. [221]</td></tr><tr><td>Cassava starch/Glycerol/ Carboxymethyl-cellulose + Turmeric oil (30 µl)</td><td>Rectangular 80 × 25 × 0.5 mm</td><td>69.47 % (2 months)</td></tr><tr><td>Plasticised chitosan</td><td>Rectangular 40 × 50 × 0.150 mm</td><td>≈ 56 % (26 days)<br>≈ 62 % (26 days)<br>≈ 78 % (26 days)</td></tr><tr><td rowspan="11">Compost rich in organic matter</td><td rowspan="11">Perforated compost bin<br>Cylindrical composter (10 L)</td><td>Plasticised chitosan/ Hydrolysed-orange peel (30/ 70; w)</td><td rowspan="3">Rectangular 0.5 mm (thickness)</td><td rowspan="11">20–55°C</td><td>≈ 34 % (9 weeks)</td><td rowspan="11">Aversa and Barletta [222]<br>Bulatović et al. [223]</td></tr><tr><td>Hydrolysed orange peel</td><td>0.05 % (16 days)</td></tr><tr><td>PBAT/TPS/PLA/Talc/Amide wax (45/25/20/8/2; w)</td><td>1.29 % (16 days)</td></tr><tr><td rowspan="4">LDPE/TPS (90/10; w)</td><td>LDPE/TPS (70/30; w)</td><td>1.53 % (16 days)</td></tr><tr><td>LDPE/TPS (60/40; w)</td><td>7.44 % (16 days)</td></tr><tr><td>LDPE/TPS (50/50; w)</td><td>13.47 % (16 days)</td></tr><tr><td>LDPE + MA-g-SEBS<sup>(e)</sup></td><td>0 % (16 days)</td></tr><tr><td>LDPE/TPS + MA-g-SEBS (90/ 10; w)</td><td>0 % (16 days)</td></tr><tr><td>LDPE/TPS + MA-g-SEBS (80/ 20; w)</td><td>1.19 % (16 days)</td></tr><tr><td>LDPE/TPS + MA-g-SEBS (70/ 30; w)</td><td>2.61 % (16 days)</td></tr><tr><td>LDPE/TPS + MA-g-SEBS (60/ 40; w)</td><td>5.73 % (16 days)</td></tr><tr><td>LDPE/TPS + MA-g-SEBS (50/ 50; w)</td><td>9.7 % (16 days)</td></tr><tr><td rowspan="4">Mature compost</td><td>TPS + SA (Salicylic acid)</td><td></td><td></td><td>85 % (75 days)</td><td rowspan="4">Dos Santos Filho [224]</td></tr><tr><td>PCL</td><td></td><td></td><td>≈ 10 % (60 days)</td></tr><tr><td>PCL-Oil (Macafá) (90/10; v)</td><td></td><td></td><td>≈ 7 % (60 days)</td></tr><tr><td>PCL-Flour-Oil (Macafá) (80/ 10/10; v)</td><td></td><td></td><td>≈ 23 % (60 days)</td></tr><tr><td rowspan="2">Biowaste/Bulking agent (1/1; w)</td><td rowspan="2">Compost bin (200 L)</td><td>PCL-Flour (Macafá) (90/10; v)</td><td></td><td></td><td>≈ 44 % (60 days)</td><td rowspan="2">Fernández-d'Arlas [225]</td></tr><tr><td>Keratin (Plastic film)</td><td>Rectangular 15 × 20 × NC mm</td><td>≈ 100 % (5 days)</td><td></td></tr><tr><td rowspan="2">Cellulose acetate</td><td>Keratin/Glycerol (72/28; w)</td><td></td><td></td><td>≈ 38 % (5 days)</td><td rowspan="2">Leppänen et al. [226]</td></tr><tr><td>Keratin/Glycerol (72/28; w) (cross-linked at 80°C)</td><td></td><td></td><td>≈ 43 % (5 days)</td></tr><tr><td>Cellulose basedplastic bag (PlasBel)</td><td>Cellulose (regenerated from ionic liquid)</td><td>Square 50 × 50 × 0.74 mm</td><td>≈ 20–70 °C</td><td>100 % (12 weeks)</td><td></td></tr><tr><td>Carboxymethyl cellulose (cross-linked with aluminum)</td><td>Rectangular 50 × 50 × 0.32 mm</td><td></td><td></td><td>100 % (12 weeks)</td><td></td></tr><tr><td>Nanocellulose</td><td>Rectangular 50 × 50 × 0.28 mm</td><td></td><td></td><td>100 % (12 weeks)</td><td></td></tr><tr><td>Methyl cellulose</td><td>Rectangular 50 × 50 × 0.30 mm</td><td></td><td></td><td>100 % (12 weeks)</td><td></td></tr><tr><td>Cellulose acetate</td><td>Rectangular 50 × 50 × 0.65 mm</td><td></td><td></td><td>0 % (12 weeks)</td><td></td></tr><tr><td>Cellophane</td><td>Rectangular 50 × 50 × 0.28 mm</td><td></td><td></td><td>100 % (12 weeks)</td><td></td></tr><tr><td>Wet strength paper</td><td>Rectangular 50 × 50 × 0.11 mm</td><td></td><td></td><td>100 % (12 weeks)</td><td></td></tr><tr><td>Cellulose carbamate</td><td>Rectangular 50 × 50 × 0.33 mm</td><td></td><td></td><td>100 % (12 weeks)</td><td></td></tr><tr><td>Butylated hemicelluloses (DS<sup>(f)</sup> :1)</td><td>Rectangular 50 × 50 × 0.83 mm</td><td></td><td></td><td>100 % (12 weeks)</td><td></td></tr><tr><td>Cellulose palmitate</td><td>50 × 50 × 0.175 mm</td><td></td><td></td><td>0 % (12 weeks)</td><td></td></tr><tr><td>Cellulose octanoate</td><td>50 × 50 × 0.129 mm</td><td></td><td></td><td>0 % (12 weeks)</td><td></td></tr></tbody></table>

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<table><caption>**Table 2 (continued)**</caption><thead><tr><th>Decomposition media</th><th>Type of composter</th><th>Materials</th><th>Shape & size of samples (l, w, t)<sup>(a)</sup></th><th>T° (min/max)</th><th>Results (Weight loss)</th><th>Authors</th></tr></thead><tbody><tr><td rowspan="10">Composted soil</td><td rowspan="10">Compost bin (6.7 L)</td><td>PLA</td><td>Rectangular 30 × 10 × 4 mm</td><td></td><td>≈ 0.3 % (45 days)<br>≈ 0.75 % (45 days)<br>≈ 1.4 % (45 days)<br>≈ 2 % (45 days)</td><td>Pokhriyal and Rakesh [227]</td></tr><tr><td>PLA/Untreated-Fiber (95/5; w)<br>(Himalayacalamus falconeri)</td><td></td><td></td><td>≈ 2.2 % (45 days)<br>≈ 0.6 % (45 days)<br>≈ 1.2 % (45 days)<br>≈ 1.6 % (45 days)</td><td></td></tr><tr><td>PLA/Untreated-Fiber (90/10; w)</td><td></td><td></td><td>≈ 2.2 % (45 days)<br>≈ 0.6 % (45 days)<br>≈ 1.2 % (45 days)<br>≈ 1.6 % (45 days)</td><td></td></tr><tr><td>PLA/Untreated-Fiber (85/15; w)</td><td></td><td></td><td>≈ 2.2 % (45 days)<br>≈ 0.6 % (45 days)<br>≈ 1.2 % (45 days)<br>≈ 1.6 % (45 days)</td><td></td></tr><tr><td>PLA/Untreated-Fiber (80/20; w)</td><td></td><td></td><td>≈ 2.2 % (45 days)<br>≈ 0.6 % (45 days)<br>≈ 1.2 % (45 days)<br>≈ 1.6 % (45 days)</td><td></td></tr><tr><td>PLA/Alkali-Treated-Fiber (95/5; w)</td><td></td><td></td><td>≈ 2.2 % (45 days)<br>≈ 0.6 % (45 days)<br>≈ 1.2 % (45 days)<br>≈ 1.6 % (45 days)</td><td></td></tr><tr><td>PLA/Alkali-Treated-Fiber (90/10; w)</td><td></td><td></td><td>≈ 2.2 % (45 days)<br>≈ 0.6 % (45 days)<br>≈ 1.2 % (45 days)<br>≈ 1.6 % (45 days)</td><td></td></tr><tr><td>PLA/Alkali-Treated-Fiber (85/15; w)</td><td></td><td></td><td>≈ 2.2 % (45 days)<br>≈ 0.6 % (45 days)<br>≈ 1.2 % (45 days)<br>≈ 1.6 % (45 days)</td><td></td></tr><tr><td>PLA/Alkali-Treated-Fiber (80/20; w)</td><td></td><td></td><td>≈ 2.2 % (45 days)<br>≈ 0.6 % (45 days)<br>≈ 1.2 % (45 days)<br>≈ 1.6 % (45 days)</td><td></td></tr><tr><td>Plastic tray (1 L)</td><td>CS<sup>(g)</sup></td><td>Square 20 × 20 × 0.076 mm</td><td></td><td>27 % (3 weeks)</td><td>Riaz et al. [228]</td></tr><tr><td rowspan="3">Composted soil</td><td rowspan="3">Plastic tray (1 L)</td><td>CS/CRE<sup>(h)</sup> (99/1; w)(Chinese chive root extract)</td><td>Square 20 × 20 × 0.085 mm</td><td></td><td>≈ 32 % (3 weeks)</td><td></td></tr><tr><td>CS/CRE (97/3; w)</td><td>Square 20 × 20 × 0.098 mm</td><td></td><td>≈ 40 % (3 weeks)</td><td></td></tr><tr><td>CS/CRE (95/5; w)</td><td>Square 20 × 20 × 0.113 mm</td><td></td><td>47.36 % (3 weeks)</td><td></td></tr><tr><td>Mature compost/Composted soil</td><td>PCL</td><td>Square 10 × 10 × 3 mm</td><td></td><td>16 % (8 weeks)</td><td>González Petit [201]</td></tr><tr><td rowspan="3">Poultry manure/Wheat straw/Tomato crop residues (1/0.2/2; w)</td><td rowspan="3">Poultry manure/Wheat straw/Tomato crop residues (1/0.2/2; w)</td><td>PCL/ES<sup>(i)</sup> (1/1; w)</td><td>Arcade support</td><td>≈ 20–55 °C</td><td>≈ 27 % (16 weeks)</td><td>Malińska et al. [200]</td></tr><tr><td>GS2189<sup>(j)</sup></td><td>Rod support</td><td></td><td>≈ 32 % (16 weeks)</td><td></td></tr><tr><td>GS2189/Biochar (95/5; w)</td><td>Arcade support clip</td><td></td><td>≈ 20 % (16 weeks)</td><td></td></tr><tr><td rowspan="2">Mature compost/Potato/Wood chips/ Household waste/Fresh grass/Rabbit droppings (0.36/0.22/0.17/0.15/ 0.7/0.3; w)</td><td rowspan="2">Drum composter (70 L)</td><td rowspan="2">PHBV (HV 5 %)</td><td rowspan="2">Square 50 × 50 × 0.350 mm</td><td rowspan="2">7.5°C - 39.1°C</td><td>≈ 19 % (16 weeks)</td><td rowspan="2">Lyshtva et al. [229]</td></tr><tr><td>8.74 % (8 weeks)</td></tr></tbody></table>

to allow for a comparative analysis of the biodegradability and/or disintegration performance of different materials.

Additional relevant data were compiled (Tables 1 and 2) to facilitate a more in-depth comparative analysis and to identify key variables contributing to the variability in biodegradation and disintegration rates observed across studies.

(i) A precise identification of the polymer type and composition was performed, including any blends or additives, to evaluate the relative influence of environmental (biotic/abiotic) factors and the impact of formulation components on biodegradability. Some commercially available formulations were also included, to reflect the actual performance of biodegradable consumer products under home composting conditions.

(ii) The most detailed possible description of composting conditions was compiled, when available, to ensure a fair comparison between experiments. Across all studies, the collected data included: total test duration, the shape and dimensions of the samples, and the exact composition of the decomposition medium.

More specifically, for laboratory-scale experiments, additional information was gathered—when reported—on the reactor temperature, as well as structural parameters characterizing the evolution of the samples during biodegradation, such as the weight-average molecular mass ($M_w$, $M_n$), polydispersity index (PI), and crystallinity degree ($\chi$ %).

For real-life home composting studies, the following information was also recorded: type of composter used, volume of the composting medium, and the minimum and maximum temperatures recorded throughout the study period.

These parameters—particularly temperature, medium composition, and sample shape and size—are likely to explain the variations in biodegradability observed among materials with similar or even identical chemical structures.

This structured comparison allowed for the identification of key variables influencing performance and revealed certain shortcomings in the reproducibility and standardization of experiments, particularly under real-life home composting conditions.

The following sections list the main results of published studies on the decentralized compostability of biodegradable plastics, conducted under laboratory conditions (Table 1), and under uncontrolled conditions (Table 2).

# 7. Factors influencing the compostability of biodegradable plastics

## 7.1. Shape and dimensions of the fragments

Enzymatic hydrolysis is strictly a surface degradation mechanism, as it requires direct contact between extracellular enzymes and macromolecular polymer chains. In the absence of cavities, cracks, or microchannels, microbial enzymes are too large to penetrate the interior of the
fragment. While non-enzymatic chemical hydrolysis occurs simulta-
neously at the surface and at the core of the sample in a variable relative
proportion, and which depends mainly on two parameters (i.e. diffu-
sivity of water in the fragment, and thickness of this) [230]. In
non-enzymatic chemical hydrolysis, once a polymer fragment reaches a
critical thickness, bulk degradation becomes the predominant abiotic
mechanism. The reaction then becomes autocatalyzed due to the accu-
mulation of hydrolysis by-products (e.g. H₃O⁺ or OH⁻) depending on
whether the environment is acidic or alkaline [231].

The shape, as well as the dimensions of the plastic material play an
important role in the speed of the degradation process under mesophilic
composting conditions [148,232–234]. Indeed, the specific surface area
(i.e. surface/volume ratio) of a plastic fragment greatly impacts its
degradation rate (Fig. 3). A higher specific surface area relative to the
volume of plastic fragments increases their exposure to hydrolytic en-
zymes involved in biological degradation, leading to a faster reduction
in molecular weight. In addition, the surface roughness of plastic frag-
ments is also an important parameter in their biodegradability. Frag-
ments with greater surface roughness present a larger area exposed to
extracellular hydrolytic enzymes and offer a more favourable niche for
microorganisms, thereby facilitating their attachment and enrichment
on the material through biofilm formation (Fig. 3) [35].

In this context, Jandas et al. [235] demonstrated that rectangular
PLA fragments with a thickness of 30 µm reached the 90 % minerali-
zation threshold in approximately 60 days in a vermicompost-type
medium at temperatures below 30 °C. In contrast, a film thickness of
200 µm was insufficient to achieve the same result, whether at 25 °C or
37 °C [134]. It was also shown that a polymer blend composed of
PLA/PHB/TPS/plasticizer, in the form of 5 × 5 mm square fragments

with a thickness of 40 µm, reached the 90 % biodegradation threshold
more rapidly—between 82 and 100 days depending on the relative
proportions of the constituents in the tested formulations—compared to
thicker fragments (350 µm), which only surpassed this standard after
140 days [206].

Polymeric materials in powder form exhibit a much higher surface area-to-volume ratio than film samples, significantly enhancing their exposure to extracellular hydrolytic enzymes present in the medium. Adhikari et al. [196] demonstrated that samples of PBS/starch powder, PBS, and PLA exhibited biodegradation rates of 24.4 %, 16.8 %, and 13.8 %, respectively, after 28 days of soil burial. In contrast, the same polymers in the form of plastic films with a thickness of 2 mm showed markedly lower biodegradation rates of 7.2 %, 1.2 %, and 0 %, respectively, under identical conditions.

7.2. *Bio-augmentation and biostimulation*

The concepts of bioaugmentation and biostimulation stem from the
broader notion of bioremediation, which involves using microorganisms
to detoxify or remove unwanted waste from contaminated sites
[236–239]. Bioaugmentation is a method of applying autochthonous or
non-native microorganisms, wild-type or genetically modified, to
waste-contaminated sites to accelerate their disposal [240]. This
approach can involve one or several isolated microbial strains, or entire
microbial communities, commonly introduced in the form of mature
compost extracts [241].

While biostimulation refers to the addition of nutrients—broadly
defined as substances that limit microbial growth or enhance enzymatic
activity—such as phosphorus, nitrogen, or enzymatic cofactors (e.g.

<figure><img src="image_4.png" /><figcaption>**Fig. 3.** Schematic illustration of the influence of size, shape and surface properties on the biodegradation process. (A): The formation of a biofilm on the surface of a plastic product involves microbial adhesion and enrichment. They depend on the interaction between secreted microbial extracellular polymeric substances (EPSs) and the surface properties of the plastic material (roughness, surface energy, etc.). (B): The thickness and diffusivity of water within the matrix of the plastic fragment are the two main parameters governing the non-enzymatic chemical hydrolysis of macromolecular chains (surface and bulk degradation). (C): The fragmentation capacity of a plastic product modulates its surface/volume ratio during the biodegradation process. A higher surface/volume ratio means greater exposure of macromolecular chains to extracellular microbial hydrolytic enzymes, leading to more rapid depolymerisation. *L*: sample thickness; *Lcrit*: critical sample thickness; *D*: water diffusion coefficient; *λ*': pseudo first order rate of hydrolysis; *Ys*: surface energy of the plastic fragment; *Ys*<sup>d</sup>: dispersive component of surface energy; *Ys*<sup>p</sup>: polar component of surface energy. Created in BioRender. Bou sarkis, A. (2024) BioRender.com/d40z037.</figcaption></figure>
divalent metals), its goal is to stimulate the indigenous microorganisms. This stimulation aims to enhance their ability to biodegrade pollutants more effectively [113].

The addition of microbial inocula—both bacterial and fungal—has been shown to significantly enhance the biodegradability of certain polymers under mesophilic conditions. This approach has enabled the effective degradation of polymers such as PLA, which are otherwise considered recalcitrant in such environments [241,242]. This effect is attributed to the conformational specificity of extracellular hydrolytic enzymes (e.g., esterases, proteases, lipases) secreted into the environment, which enhances their efficiency in cleaving the ester bonds of high molecular weight polymers [113]. This approach could help render such polymers compostable at home by introducing the specific, non-ubiquitous microbial flora required for their breakdown.

For example, Mistry et al. [113] demonstrated a substantial enhancement in the biodegradation of PLA fragments under mesophilic composting conditions, following the addition of a bacterial inoculum composed of strains specifically selected for their PLA-degrading capabilities. This improvement was also associated with a notable reduction in the phytotoxicity of the resulting compost.

For their part, Jeon and Kim [145] highlighted that the use of sterilized compost inoculated with a single microbial strain capable of efficiently degrading PLA under mesophilic conditions significantly increased the biodegradation rate of PLA fragments in the medium. A degradation rate of 31 % was achieved within 40 days, in contrast to the typically observed rates for this polyester under similar conditions, which are generally below 5 %.

Microbial taxa can also be introduced as a community, typically through the incorporation of mature compost into a biowaste pile. In this context, Wang et al. [243] demonstrated that the addition of mature compost (20 %, w/w) to a pile of decomposing biowaste (composed of food waste, dried blood, and corn cobs in a 5:1:1 w/w ratio), containing 5 % PBAT, in 8 L composters supplied with an air flow heated to 50 °C, resulted in a 16.7 % increase in the disintegration rate of polymer fragments compared to a control treated with sterilized mature compost. This addition significantly altered the microbial community at the genus level over time. Three genera—*Thermobifida*, *Garciella*, and *Bacillus*— which include functionally beneficial strains for the composting process, were overrepresented in the non-sterilized treatment compared to the control. Notably, *Bacillus* taxa have been previously reported to effectively depolymerize PBAT fragments via the secretion of hydrolases [244–246]. Furthermore, the authors observed a 25.91 % reduction in the global warming potential in composters enriched with mature, non-sterilized compost, attributed to lower CH₄ and N₂O emissions. The incorporation of mature compost thus plays a dual role: (i) it acts as a bulking agent, enhancing porosity and overall aeration of the biowaste pile; and (ii) it enriches the environment with aerobic microbial strains, which outcompete anaerobic bacteria, thereby limiting CH₄ production [247].

In addition to directly enhancing the biodegradation of various polymers through the action of specific enzymes, bio-augmentation using microbial agents can also help reduce environmental acidification during the composting process. This, in turn, generally promotes the decomposition of organic matter and may indirectly improve the compostability of biodegradable plastics under decentralized composting conditions [248]. Indeed, household post-consumer food biowaste contains substantial amounts of sugars and lipids, which, during the composting cycle, contribute to significant acidification of the environment. This acidification can partially inhibit microbial activity, thereby delaying the degradation of organic matter and, in some cases, leading to the failure of the composting process [249,250]. In this regard, a group of bacteria specialized in the biodegradation of organic acids has been shown to significantly increase the pH of a composting pile composed of food waste and maize haulms (75/25, w/w). This effect was observed when the bacteria were introduced exogenously into the decomposition medium. At the beginning of the trial, the pH was

relatively low (5.77), but it rose to 8.67 after seven days and remained relatively stable over the following three weeks. This pH increase was accompanied by a rise in the internal temperature of the pile and a reduction in the phytotoxicity of the resulting compost [251].

Thus, bioaugmentation represents an innovative and promising strategy to overcome the limited biodegradability of certain polymers under mesophilic conditions. This can be achieved either directly—through the cleavage of macromolecular chains via the extracellular secretion of specific enzymes—or indirectly, by improving the overall performance of the composting process.

Regarding the biostimulation strategy, Mayekar and Auras [252] tested several approaches to stimulate microbial communities, primarily to trigger the biotic degradation of their PLA samples under mesophilic conditions. These strategies were implemented during a mineralization experiment conducted at a constant temperature of 37 °C. They incorporated iron oxide (Fe₃O₄) in nano-powder form, skim milk, gelatine, and ethyl lactate into the biodegradation substrate (i.e. mature compost/vermiculite). These additives were supposed to stimulate different steps in the PLA biodegradation process: (i) trigger the proteolytic activity of microorganisms (i.e. gelatin and skim milk); (ii) increase the efficiency of microorganisms to specifically biodegrade PLA (i.e. ethyl lactate); and (iii) act as a catalyst to increase the rate of non-enzymatic chemical hydrolysis reactions (Fe₃O₄). It appears that, except for iron oxide, each bio-stimulant greatly increases the rate of PLA mineralization: gelatine (56.4 % in 180 days); skimmed milk (87.7 % in 180 days); ethyl lactate (105 % in 180 days, activation effect); compared to PLA alone, which has a negative mineralization rate (~30 % in 180 days). The increase in the rate of mineralization of biostimulated reactors is accompanied by a greater decrease in molecular weight (Mn), and a greater increase in the crystallinity rate of samples compared to those whose degradation medium has not been biostimulated. As far as iron oxide is concerned, despite the observation of negative mineralization rates (~15 % alone, and ~150 % when combined with gelatine), its use alone or in combination, significantly decreases the molecular weight of PLA samples over time, in addition to increasing their crystallinity rate (28.3 % at t0, 31.4–40.9 % at t + 180). These encouraging findings suggest that the external addition of small amounts of biostimulants to the composting medium may enable high molecular weight PLA-based plastics to degrade effectively under home composting conditions.

Biostimulants applied individually can each independently enhance the biodegradability of bioplastics under mesophilic conditions. However, these approaches may also be combined to potentially act synergistically and maximize the breakdown of biodegradable plastics under decentralized composting conditions.

In this regard, a separate study analogous to the first, conducted by Mayekar and Auras [210], provides compelling evidence for the potential synergy between organic biostimulants and inorganic additives under mesophilic composting conditions. The treatment combining gelatin and Fe₃O₄ led to the most significant acceleration in degradation, reducing the time to reach a number-average molecular weight (Mn) ≤ 10 kDa by 30 % compared to the untreated control. While gelatin alone resulted in a 25 % reduction and Fe₃O₄ achieved a 17 % improvement, their combined use produced a greater effect than either component individually, clearly indicating a synergistic interaction. The authors attribute this to the dual role of gelatin as an enzyme inducer, stimulating microbial protease activity, and Fe₃O₄ as a hydrolytic catalyst that facilitates chain scission through redox mechanisms. These findings demonstrate that the targeted co-application of enzymatically active substrates and catalytic additives can effectively overcome the thermal limitations of PLA biodegradation under mesophilic conditions, offering a promising strategy for improving home-composting performance.

### 7.3. Pre-treatments

There are several methods for pre-treating plastic materials, applied before they go through an organic recovery route (e.g. composting),
which aim to accelerate their biodegradation [212,253]. They act upstream by an initial deterioration of the plastic material, to make it more likely to be biodegraded by microorganisms afterwards. Non-thermal physical pre-treatment methods are potentially less energy-consuming, and less polluting compared to physical, thermal and chemical methods. They include treatments: (i) UV; (ii) by high-frequency ultrasound; (iii) mechanical grinding and (iv) the use of plasma. UV and high-frequency ultrasonic treatment act by breaking the chemical bond, leading to a decrease in the molecular weight of the polymer, and even to chemical modifications [254]. The decrease in the molecular weight of polymers makes them more likely to be degraded by hydrolysis, as smaller molecules increase water diffusion and the accessibility of hydrolytic extracellular enzymes to macromolecular chains [255,256]. While changing the surface chemical composition of plastic films can modulate the affinity of microorganisms to adhere it and thus increase their capacity to form a biofilm [253,257]. Oxygen enrichment of the plastic product allows for better adhesion of many microorganisms and thus better biodegradation of plastic materials and must be sought [138].

Under mesophilic composting conditions at a constant temperature of 37 °C, the use of UV irradiation, and the application of high-frequency ultrasound, do not seem to increase the biodegradability of PLA, which remains recalcitrant at this temperature [138].

This could be attributed to the limited extent of chain scission or oxidation caused by UV or ultrasound treatment, which may not be sufficient to lower the molecular weight below the critical threshold (approximately 10,000 to 20,000 g/mol) necessary for efficient biodegradation and mineralization of PLA [258].

Indeed, under mesophilic composting conditions, the primary limitation to PLA biodegradation lies in its glassy physical state ($T < T_g$), which significantly restricts the accessibility of ester bonds to hydrolytic cleavage [259]. This severely limits the molecular weight reduction required to generate low-molecular-weight oligomers that can be assimilated by microorganisms

A polymer with an initially lower molecular weight is more prone to undergo abiotic hydrolysis, as shorter chains are more accessible to water molecules and more susceptible to backbone cleavage, thereby further accelerating the decrease in molecular weight until the threshold is reached [260]. However, unlike thermophilic composting conditions where abiotic (non-enzymatic) hydrolysis is significant, the rate of chemical hydrolysis under mesophilic conditions remains very low [261]. As a result, the initial effect of the pre-treatment is insufficient to trigger substantial molecular weight reduction, thereby failing to initiate PLA mineralization.

Nevertheless, when compost is first sterilized and then inoculated with an efficient strain for PLA biodegradation (i.e. a bioaugmentation strategy), UV exposure has been shown to enhance the biodegradability of the polymer under mesophilic conditions at 37 °C [145]. The application of UV radiation has also been shown to significantly enhance the biodegradability of other biopolymers under controlled home-composting conditions (25–30 °C), particularly poly (3-hydroxybutyrate) (PHB) and a copolymer consisting of PCL urethane and PHB [262]. The biodegradation rate was found to increase with UV exposure time—at a wavelength of 254 nm—up to the maximum tested duration of 9 h. This effect is attributed to chain scission and photo-oxidation mechanisms, which enhance the susceptibility of polymers to microbial degradation under mesophilic conditions. Differential Scanning Calorimetry (DSC) revealed only minor changes in the crystallinity of PCL and PHB segments in the copolymer, suggesting that photochemical reactions primarily occur in the amorphous regions.

### 7.4. Polymer blends and copolymer composition

Manny biodegradable plastics suffer, compared to their non-biodegradable counterparts, from lower mechanical properties, such as flexibility, strength and toughness, which has the effect of restricting

their use [123]. To overcome this limitation, polymer blending can be an attractive solution with the intention of producing plastic materials with sufficient mechanical properties to be able to perform certain functions, such as packaging, while reducing the overall cost of mixing [207,263]. The polymer mixture results in a wide variety of complex phase structures, ranging from co-crystallization to immiscible phase separation that impacts the mechanical, thermal and biodegradative properties [207,263,264]. In particular, it has been demonstrated that a polyester blend composed of the recalcitrant polymer PLA and the home-compostable polymer PCL (PLA/PCL, 80/20; w/w) exhibits a significantly higher biodegradability under home composting conditions (>90 % in 265 days) compared to PLA alone. In addition to improve its mechanical properties. Conversely, a formulation composed of PLA and PHB in the same proportions fails to reach the 90 % mineralization threshold required for home compostability certification [123]. Fogašová et al. [206] reported a biodegradation rate of approximately 90 % after 140 days for a formulation composed of PLA/PHB/citric acid esters (plasticizer) in a 70/15/15 (w/w) ratio. According to the authors, the presence of the plasticizer likely promotes a re-esterification process between PLA and PHB segments. This reaction, combined with the higher biodegradative potential of PHB, facilitates the breakdown of the copolyester into lower molecular weight fragments, thereby enhancing the overall biodegradation of the material [265]

More recently, Han et al. [213] reported significant biodegradability (28.7 % to 82.7 % mineralization within 45 days) of poly (L-lactic acid) (PLLA) fragments blended with amorphous PHA (PH34B, containing 33 mol % H4B; aPHA) in varying proportions (9:1, 7:3, 5:5; w/w), under controlled laboratory conditions at 30 °C. This enhanced biodegradability is primarily attributed to a reduction in the crystallinity of PLLA fragments caused by the anti-nucleating effect of aPHA during the extrusion step, which also increases their hydrophilicity.

Moreover, atomic force microscopy (AFM) revealed a pronounced phase separation between the two polymers, without aggregation. The faster degradation of the aPHA phase leads to the formation of microfluidic channels that facilitate the diffusion of microbial hydrolytic enzymes to the polymer interface. This, in turn, increases the surface area of PLLA exposed to enzymatic attack, thereby accelerating its biodegradation.

PHBV is a copolymer of 3-hydroxybutyrate and 3-hydroxyvalerate (3-HV), and within the PHA family, it holds the greatest potential as a substitute for petroleum-based polymers in the packaging sector. Its mechanical properties are comparable to those of polypropylene and polystyrene in terms of strength and flexibility. Moreover, its gas barrier performance is close to that of PET, while meeting regulatory standards for food contact applications [209,266,267].

In addition, PHBV exhibits improved biodegradability compared to PHB [268]. The rate of biodegradation depends on the relative proportion of its monomer units. A 3-HV content of 12 % has been shown to promote greater biodegradability than higher proportions (43 %, 47 %, 52 %, 64 %, 72 %) in a 15.9-week mesophilic test using a mature soil-compost medium [269]. Conversely, Mergaert et al. [197] reported over tenfold higher weight loss for a copolymer containing 20 % 3-HV compared to one with 10 %

For their part, under real home composting conditions, Lyshtva et al. [216] observed 8.78 % disintegration of PHBV (containing 5 % HV) film fragments after 8 weeks. The absence of significant changes in surface chemical composition (as revealed by FTIR) and thermal properties (as shown by DSC) suggests that the degradation mechanism was limited either to simple mechanical fragmentation and/or to surface enzymatic hydrolysis, without altering the overall chemical structure of the polymer.

Polymer blending is also a strategy used to reduce the cost of polymer resins in the manufacture of plastic products—particularly when a polymer's high price limits its application in sectors such as food packaging, as is the case for poly(ε-caprolactone) (PCL), which costs
approximately €14/kg. Aragón- Aragón-Gutiérrez et al. [229] produced plastic fragments by thermo-extrusion of PCL blended with thermo-plasticized gliadins (TPGA) at 30 % and 50 % (w/w). They observed a marked increase in mineralization rates at 25 ± 5 °C for the blends compared to neat PCL (PCL: ~33 %; PCL/TPGA 70/30: 79 %; PCL/TPGA 50/50: 86 %).

This behaviour was attributed to a significant enhancement in water adsorption capacity (PCL: 0.05 %; PCL/TPGA 70/30: 8.1 %; PCL/TPGA 50/50: 16.3 %) and an increase in surface roughness of the blends, both of which contribute to improved microbial accessibility and degradation.

## 7.5. Biodegradable composites

A biodegradable composite material is composed of at least one polymer matrix and one or more immiscible fillers. A wide variety of organic substances can be used as fillers in such materials, including natural fibers and agro-industrial waste. The filler is typically present in smaller proportions than the matrix to avoid excessive deterioration of the composite's mechanical properties. Fillers are introduced to: (i) reduce production costs; (ii) enhance material properties (e.g., mechanical, thermal, gas barrier performance, biodegradability); (iii) adjust the density of the final product; and (iv) minimize environmental impact [211,270]. The incorporation of low-value fillers in composite manufacturing is a promising strategy for promoting the development of biobased plastics, particularly given their relatively high production cost [271].

### 7.5.1. Biodegradable composites reinforced with natural fibres

Several natural fibers can be combined with polymer matrices to produce reinforced composites [272,273]. They have several advantages over inorganic fibers, such as lower cost, lower density, biodegradability, renewable origin, and greater recyclability [211,274]. They are mainly composed of cellulose, hemicellulose and lignin, which gives them a highly hydrophilic character [275,276]. Currently, the bast fibers most used are flax, hemp, kenaf, jute, and lyocell fibers.

Composite materials including natural fibers biodegrade significantly better compared to polymer alone in real or laboratory composting conditions [215,218,219,277]. The nature of the fibers (e.g. jute, hemp, lyocell, bamboo), their proportion within the composite, orientation (unidirectional vs. multidirectional), average length, and the application of any pretreatment are considered key parameters. These characteristics have been identified as major factors influencing both the biodegradability of composites under home composting conditions and their mechanical performance [227].

### 7.5.2. Effects of natural fiber pretreatments on composite biodegradability

The excessive presence of undesirable constituents on the surface of natural fibers (e.g. waxes, hemicellulose, lignin, pectin) can hinder the formation of interfacial bonds with the polymer matrix, resulting in poor adhesion and, consequently, reduced mechanical performance of the composite material [278,279]. To address this issue, various chemical surface treatments—such as alkaline, acid, acetylation, benzoylation, peroxide, or silane treatments—can be applied to modify fiber surfaces and improve compatibility with the matrix [280].

However, such modifications often decrease the biodegradability of the composite, as they reduce the hydrophilic nature of the fibers. This trade-off, while enhancing interfacial compatibility with the typically hydrophobic matrix, may negatively impact the material's end-of-life environmental performance [281].

Pokhriyal and Kumar Rakesh [219] observed that applying an alkaline pretreatment (5 % w/v NaOH) to natural fibers derived from the red bamboo plant (Himalayacalamus falconeri) effectively removed surface impurities. This treatment enhanced the tensile, flexural, and impact properties of the resulting composite. Nevertheless, it led to a reduction in the weight loss of the treated composite under home

composting conditions, indicating a lower degree of disintegration.

#### 7.5.2.1. Synergistic effects of the presence of fiber on matrix bio-degradation
Liu et al. [282] reported a strong synergistic effect of filler incorporation on the disintegration of a composite based on poly (butylene succinate) (PBS) and jute fibers (90/10, w/w), tested in a composted soil-like medium under mesophilic conditions. After 180 days, the composite exhibited a weight loss of 62.5 %, which was significantly higher than that of PBS alone (31.4 %) and jute fibers alone (24.7 %).

Interestingly, no positive linear relationship was observed between the extent of weight loss and the amount of filler incorporated within the composite across the tested proportions (10 %, 20 %, 30 %). On the contrary, a decrease in disintegration was noted for filler contents above 10 %, with weight loss decreasing progressively from 20 % to 30 % fiber content.

This accelerating effect of the filler on polymer matrix disintegration was not observed in the case of a PLA/non-woven flax fiber composite, tested under controlled mesophilic composting conditions (25 ± 2 °C; 40–50 % humidity) [136]. In that case, the increased weight loss of the composite compared to neat PLA appeared to result mainly from the disintegration of the flax fibers, with no significant effect on the degradation of the polymer matrix itself, despite the use of high fiber contents (40 %, 50 %, 60 %; w/w).

#### 7.5.2.2. Fiber orientation
Under real composting conditions, a composite based on PLA and multidirectional flax fibers (50/50, w/w) was shown to disintegrate more rapidly than an equivalent composite containing unidirectional flax fibers, with respective disintegration rates of 27 % and 19 % [277]. This difference was primarily attributed to the enhanced water absorption capacity of the non-woven (i.e., multidirectional) fiber-based composite (~40 % vs. ~10 %), resulting from its higher porosity (18 % vs. 4 %) and lower crystallinity of the polymer matrix (34 % vs. 42 %). These characteristics imply increased material brittleness due to swelling and shrinkage induced by water absorption.

#### 7.5.2.3. Impacts of the length and dispersion of natural fibers
Gunning et al. [215] investigated the effect of incorporating fibers from different plant species (i.e., hemp, jute, and lyocell) on the disintegration rate of PHB-based composites. Their results showed that composites reinforced with jute fibers degraded more rapidly than those containing hemp or lyocell. This enhanced disintegration was attributed to the higher water sorption capacity and improved dispersion of jute fibers within the polymer matrix.

The authors highlighted the shorter average length of jute fibers compared to hemp and lyocell, as well as a greater number of hydrogen bonds (as revealed by FTIR analysis), which likely contribute to increased water absorption—partly due to higher porosity at the fiber-matrix interface. Subsequently, following the degradation of the filler, this may promote enhanced hydrolytic microbial activity on the polymer matrix, owing to a larger surface area exposed to extracellular enzymes [199].

However, the short average length of jute fibers was not sufficient to improve the tensile properties of the composite. As a result, the tensile strength of the composite reinforced with 30 % jute fibers was slightly lower than that of the neat polymer (~16 MPa vs. ~18 MPa). This emphasizes the importance of the critical fiber length, which governs the efficiency of stress transfer between the matrix and the reinforcement. The aspect ratio (S = L/D, where L and D represent the fiber's length and diameter, respectively) must exceed a certain threshold—referred to as the critical aspect ratio—for fiber incorporation to effectively enhance the mechanical performance of the composite [283,284].
### 7.5.3. Biodegradable composites based on agro-industrial and household waste

Household and agro-industrial bio-waste, commonly referred to as agricultural waste (AW), can be used as a filler in combination with biodegradable polymers to produce composite materials. This type of waste is often landfilled, and its uncontrolled decomposition can lead to serious environmental issues, such as air and water pollution. Valorising AW in composite production therefore contributes to the development of environmentally friendly materials at a lower cost, fully aligned with the principles of a circular bioeconomy [285,286].

Typical examples of AW include fruit pomace, seed husks, fruit and vegetable peels, or low-grade whole biomass that cannot be used for food purposes [186,287]. In addition to offering significant cost reductions, the incorporation of AW into biocomposites can enhance several key properties of plastic materials—such as biodegradability, antioxidant and antimicrobial activity, and gas barrier performance—which are particularly valued in the packaging sector [288,289].

For example, a PCL-based composite containing Macaíba flour as a filler (90/10, w/w) exhibited significantly higher biodegradability (~44 % after 60 days) compared to neat PCL (~10 %) and a composite containing Macaíba oil (~7 %) under the same conditions [290]. The incorporation of the flour notably enhanced the biodegradability of the polymer matrix. However, this came at the expense of mechanical performance, as the elongation at break of the composite decreased significantly compared to that of the neat polymer. In contrast, the presence of Macaíba oil did not markedly affect the biodegradation rate, elongation, or modulus of the composite, suggesting a plasticizing effect of the oil.

Altieri et al. [186] reported contrasting biodegradability behaviors of PBSA-based composites filled with collagen hydrolysates obtained by either alkaline or enzymatic hydrolysis, depending on the composting conditions. Under industrial composting conditions (58 °C), the enzymatically derived collagen hydrolysate significantly increased the biodegradability of the composite (PBSA/hydrolysate: 80/20, w/w) compared to neat PBSA (87.6 % vs. 69.7 % after 20 days). However, under mesophilic conditions (25 °C), its biodegradability decreased (19.7 % vs. 46.9 % after 20 days).

This trend was not observed for composites containing collagen hydrolysate obtained by alkaline hydrolysis, which showed higher biodegradability than the polymer alone under both composting conditions. These differences were attributed to structural variations in the hydrolysates. Altieri et al. [186] demonstrated that alkaline treatment resulted in greater structural disruption of the collagen hydrolysate, as evidenced by a reduction in β-sheet content and an increase in free amino acids, making the material more susceptible to microbial degradation.

The fertilizing potential resulting from the decomposition of biodegradable plastic materials in compost-like environments is generally considered low for higher plants. This is mainly because the elemental composition of most biodegradable polymers is limited to carbon, hydrogen, and oxygen—elements that do not contribute directly to plant mineral nutrition. However, the incorporation of agricultural waste (AW) rich in nitrogen, such as by-products from the animal, dairy, or brewing industries, may enhance the fertilizing value of these materials by introducing nutrients beneficial to plant growth [224].

### 7.6. Natural additives extracted from plant biomass

Natural compounds extracted from plant biomass can be incorporated as additives in plastic materials to confer specific functionalities to the final products, such as antioxidant activity, thermal stabilization during processing, or food spoilage indication. These bio-based additives offer the advantage of reducing the environmental footprint of the materials compared to conventional petrochemical-derived compounds [291,292]. Such substances can be sourced from agricultural waste (AW), including phenolic compounds extracted from fruit and vegetable peels [293,294].

However, the bacteriostatic activity of many of these plant-derived compounds may negatively affect the biodegradability of the composite, thereby reducing its overall compostability [295,296]. For instance, it has been shown that certain phenolic compounds—such as β-carotene, juglone, morin, and curcumin—can slow down the biodegradation process under mesophilic composting conditions when incorporated into PLA or P(3,4HB)-based formulations [137]. Nevertheless, the inhibitory effect remains weak or even negligible for some compounds, such as gallic acid and quercetin blended with PHBV, which do not prevent the functionalized material from reaching the 90 % mineralization threshold required for home compostability certification (NF T51–800) [297].

## 8. Discussion of home compostability assessment methodologies

### 8.1. Experiments under controlled laboratory conditions

Under aerobic conditions, the biodegradation of organic materials results in the release of most of their carbon content as carbon dioxide (CO₂), due to microbial metabolic activity [195]. Currently, the most reliable and practical method for assessing biodegradability is based on monitoring the emission of this microbial metabolite over time, which serves as a proxy for the material's mineralization rate [47]. Accordingly, the mineralization rate is generally considered indicative of the biodegradation rate of the tested samples. However, at the end of a mineralization test, a variable fraction of the degraded organic carbon does not fully mineralize. Instead, it may persist in other forms: (i) assimilated into microbial biomass; (ii) retained in the medium as intermediate degradation products (e.g., monomers and oligomers); or (iii) transformed into humic substances [298,299].

Several approaches have been developed—mainly under thermophilic conditions—to quantify two specific carbon compartments: microbial biomass and degradation by-products. These methods are generally based on establishing a carbon balance through direct or indirect estimations, performed in an inert solid medium inoculated with a compost extract (e.g. vermiculite) [299–302]. In such setups, the plastic sample serves as the sole carbon source for microbial activity, thereby eliminating positive or negative priming effects—i.e. the stimulation or inhibition of microbial use of other carbon sources present in the medium. A more precise, though technically challenging, approach involves conducting a carbon balance in a heterogeneous solid matrix (e.g. soil or compost) by using isotope labelling (¹³C or ¹⁴C) of the sample's organic carbon to trace its transformation during biodegradation (Fig. 4) [303,304].

In this regard, some authors [298,300] reported that under thermophilic conditions, up to 26 % of the total carbon at the end of the experiment — once mineralization kinetics had reached a stationary phase — originated from biodegradable plastic fragments (i.e. PLA/TPS and TPS) and was incorporated into stable microbial biomass. A similar approach, involving a carbon balance conducted on an inert activated medium, should be applied under mesophilic conditions to a broader range of polymers. The objective would be to confirm or challenge this order of magnitude and to identify potential sources of variation, such as polymer type.

If these findings prove accurate under mesophilic conditions, this will strongly indicate that current standards for home compostability (e. g. NF T51–800) require revision. These standards should be updated to account for the multiple carbon fates of the material, which are currently overlooked by conventional mineralization tests that focus exclusively on CO₂ release.

In addition, when compared to real conditions, biodegradation studies conducted in controlled laboratory settings often overestimate the biodegradability of the materials investigated [49]. This is mainly due to the maintenance of a higher temperature and humidity of the substrate than those found in real conditions. This implies a higher average oxidative and hydrolytic degradation activity, but also a longer
<figure><img src="image_5.png" /><figcaption>**Fig. 4.** Schematic illustration of two methods for estimating the ultimate aerobic biodegradation rate of biodegradable plastics under laboratory conditions. The first is made with mature compost, the other on an inert substrate (e.g. vermiculite) activated. (A): Preparation of the decomposition medium. For the other with an inert substrate, a mineral salt solution and compost extract must be mixed with it. Isolated strains or microbial communities can be also added (Bioaugmentation). (B): Home composting assessment phase. Plastic samples are incubated in the decomposition medium in bioreactors at temperatures below 30 °C, for up to one year. Specification standards for home compostability of biodegradable plastics require quantification of the amount of carbon in samples converted to CO₂ by microbial activity. A carbon balance can be carried out to provide a more accurate estimate of the net rate of ultimate aerobic biodegradation, and carbon compartmentalization. C<sub>s</sub>: sample carbon; C<sub>g</sub>: mineralized carbon (converted to CO<sub>2</sub>); C<sub>b</sub>: carbon converted into microbial biomass; C<sub>g</sub>: carbon from water-soluble degradation by-products; C<sub>r</sub>: non-degraded residual carbon. Created in BioRender. Bou sarkis, A. (2024) BioRender.com/102m877.</figcaption></figure>

test duration that does not always reflect the actual composting times observed in domestic settings [305–307].

However, the relevance of compostability tests conducted under laboratory conditions is not to be questioned, with a view to factorial decomposition of the variability associated with the biodegradation process of plastic materials. Although the extrapolation of their results remains limited [308]. Thus, they should be supplemented like agro-nomic experimental devices, by *system trials*, which are closer to the reality on the ground.

Analytical factor experiments, typically conducted under laboratory conditions, are designed to test hypotheses about processes that are not yet fully understood. They make it possible to study, either in isolation or in limited combinations, the key parameters involved in the biodegradation of plastic materials [309]. One or more explanatory variables are introduced into a controlled experimental design—keeping all other conditions constant—to evaluate their individual or combined influence on a specific response variable, such as the mineralization rate.

By contrast, system experiments take a holistic approach. They aim to assess how well a real-world system performs across a range of diverse criteria—environmental, socio-economic, and technical—while embracing the complexity of the system, rather than reducing it to a set of isolated factors. In the context of decentralized composting (e.g. household or community composting), system experiments can serve multiple purposes: (i) test the technical feasibility of the composting setup and the coherence of the decisions made; (ii) evaluate how effectively the system meets defined objectives for a given biodegradable plastic (e.g. environmental performance, social benefits, economic outcomes); (iii) assess the efficiency of the system by comparing the resources used to the results obtained; and (iv) serve as educational tools to promote knowledge-sharing and hands-on learning [310,311].

Finally, despite all the attention paid to the standardization of compostability tests under controlled laboratory conditions, a certain variability remains and concerns the biodegradation medium (i.e. mature compost) which, by its heterogeneous nature, escapes rigorous standardization. In doing so, the evaluation of the biodegradability of

certain polymers, which are considered recalcitrant in mesophilic conditions, and which require a non-ubiquitous microbial flora, such as PLA, its blends and composites, is more delicate. In this case, attention must therefore be paid to the isolation and identification of the strains involved in the degradation process under mesophilic conditions, as well as to the characterization of the enzymes that degrade these polymers, and the associated metabolic pathways.

## 8.2. *Experiments in real conditions*

Only methods evaluating material disintegration—such as weight loss and surface disappearance—are currently used to estimate the compostability of biodegradable plastics under real composting conditions [312].

These methods only partially reflect the biodegradability of plastics, as the disintegration of a plastic product can lead to the release of stable nano- and microplastics, which may have detrimental effects on the environment. However, implementing respirometric monitoring under real home composting conditions is challenging, as it requires a suitable measurement system and a substantial amount of material.

This is because the amount of carbon dioxide produced by decomposing biowaste is significantly higher—particularly at the beginning of the process—compared to that from mature compost. Moreover, the volumes involved are also much larger (approximately 50–500 liters for a home composter, versus 1–3 liters for a laboratory bioreactor). As a result, the relative proportion of carbon dioxide originating from the microbial mineralization of plastic samples is small compared to the total CO₂ emitted by the decomposition of biowaste (e.g., food and green waste). This low proportion makes accurate quantification particularly challenging.

In the future, studies should focus on implementing mesophilic compostability tests using respirometric monitoring in pilot-scale composters, with fresh biowaste rather than mature compost. This would provide a more accurate assessment of the transposability of laboratory-scale results to real composting conditions and help refine regulatory
standards while better anticipating the potential environmental impacts of biodegradable plastics under domestic composting scenarios.

In addition, experiments conducted under real home composting conditions do not systematically monitor the evolution of temperature and humidity within the compost pile throughout the process [277,219, 290,201,225,228,313]. This lack of monitoring limits the comparability of results across studies. The volume of biowaste used as the decomposition medium also varies greatly (from 1 to 3000 L), leading to potentially contrasting thermal profiles. Smaller volumes are generally more influenced by ambient air temperature and tend to remain cooler, whereas larger volumes are more likely to heat up during the initial stages of decomposition, deviating more significantly from ambient conditions. Furthermore, key contextual information—such as the geographical location of the experimental setup, the season in which the process takes place, and basic details about the composting system (e.g., dimensions, type of aeration, indoor vs. outdoor placement)—is too rarely reported.

Moreover, the exact chemical composition of materials examined under real conditions is generally less well known than for those tested under controlled laboratory conditions (reported in 33 % of selected studies under real conditions vs. 7.4 % under controlled conditions) [190,265,228,200,214,222]. This lack of information mainly concerns compostability tests performed on manufactured products. In such cases, the trade name is often mentioned, but it does not provide specific details about the formulation's components or their relative proportions. This can be attributed to the confidential nature of such data, which is protected by industrial secrecy. Consequently, the results of compostability experiments conducted on commercial products can only serve as indicative evidence—either incriminating or exculpatory—regarding the product's end-of-life behaviour, without offering insights into the underlying biodegradation mechanisms [102].

Furthermore, the duration of composting varies widely between studies (ranging from 5 to 150 days), and it is not systematically indicated—explicitly or implicitly (e.g., through the presence of kinetic curves)—which phase of the biodegradation process (growth, plateau) the material has reached by the end of the test. Some authors also use the term "composted soil" without clarifying the origin of the material and occasionally refer to soil burial or compost biodegradation tests interchangeably. This introduces confusion and reflects a conceptual conflation between compost and soil, which are two distinct environments.

## 9. Complementary approaches to improve composting performance

The efficiency of the composting process depends on several interrelated physico-chemical parameters, among which pH, C/N ratio, moisture content, aeration, temperature, and particle size distribution of the substrate are considered the most influential [314,315]. In this context, the use of computer modelling approaches to optimise process efficiency while minimising environmental impact is gaining increasing attention [316,317].

Various practices aimed at improving both the performance of the composting process and the agronomic quality of the final compost have already been extensively investigated. Most of these strategies focus on (i) directly modifying the physicochemical properties of the composting environment through the addition of exogenous organic or mineral substrates [318]. When such additions primarily influence the physical structure of the substrate (e.g. porosity, aeration), they are referred to as *bulking agents* [318,319]. In contrast, when the added substrates act directly or indirectly on other parameters, they are more appropriately referred to as *additives* [320].

For example, perlite was identified as the most effective additive under home composting conditions compared to vermiculite, zeolite, and wood chips. It was found to enhance the mineralization of organic carbon in decaying matter while exhibiting the strongest retention

capacity for NH₄⁺ ions [321]. The cation exchange capacity (CEC) of a mineral additive, along with its selectivity for NH₄⁺, directly contributes to reducing nitrogen losses through volatilization and helps regulate NH₃ and NO₃⁻ concentrations within the substrate [98]. Based on these results, the authors concluded that perlite was the most promising mineral additive among the four tested for optimising the composting process.

These practices also include (ii) the inoculation of the decomposition medium with exogenous microorganisms—referred to as a bioaugmentation (or enrichment) strategy, which constitutes a specific subcategory of additives.

In addition to this microbial inoculation approach, (iii) bio-stimulation consists in adding biostimulants to the decomposition medium to enhance the metabolic activity of indigenous microbial communities, thereby improving the biodegradation of biodegradable plastics under decentralized composting conditions [322].

Both strategies—bio-augmentation and biostimulation—have already been discussed in detail in Section 7.2 *Bio-augmentation and Biostimulation*.

Although considered a composting method, vermicomposting can be viewed as a practice primarily aimed at accelerating the decomposition of organic matter and promoting the elimination of pathogenic microorganisms [323]. It involves the presence of earthworms within the decomposition medium, which contribute to organic matter degradation through several mechanisms. These include physical fragmentation and grinding into finer particles, as well as the synergistic action of specific enzymes and microorganisms facilitated by the passage of material through the earthworm's digestive tract [324,325].

PLA particles can be broken down by the combined action of intestinal microorganisms and enzymes (e.g. carboxylesterase) present in the digestive tract of earthworms [326,327]. Recent observations have shown that *Eisenia fetida* actively ingests microplastic particles of both PET and PLA. Compared to PET, PLA particles undergo more significant changes in size distribution once released into the soil. Sub-micrometric (0.1–1 µm) and nanometric (<0.1 µm) particles become predominant relative to micrometric particles (>1 µm), indicating that PLA fragments are more extensively degraded than PET fragments within the earthworm's digestive system [328].

*Zophobas atratus*, commonly referred to as a "super worm," is known for its high productivity as well as its rich protein and lipid content [329]. It can be valorised either through the extraction of bioactive compounds or by direct incorporation into animal and human diets. Super worms are increasingly considered potential superfoods due to their high levels of vitamin B12, iron, zinc, fiber, essential amino acids, omega-3 and omega-6 fatty acids, and antioxidants [330]. Jung et al. [331] investigated the capacity of *Z. atratus* to biodegrade PBS foam, used as its sole food source.

This capacity offers dual environmental benefits: (i) the potential use of PBS as a feedstock in the industrial rearing of super worms, and (ii) the organic elimination of PBS-based products through vermicomposting. Interestingly, when PBS foam was provided as the only food source, the mortality rate of *Z. atratus* after 28 days (95.2 ± 1.6 %) was comparable to that observed in a control group fed with oat bran (95.6 ± 4.4 %). Furthermore, an average weight gain of 5.1 ± 0.5 % was recorded per individual. These findings suggest the existence of a trophic interaction between *Z. atratus* and PBS foam.

In a similar context, Peng et al. [332] observed that the addition of PLA (10–30 %) to a staple diet based on oat bran increased both the survival rate and average weight of mealworms (*Tenebrio molitor*), compared to a diet composed exclusively of the reference food (i.e., oat bran). Notably, the diet containing 20 % PLA resulted in the highest weight gain among mealworms (98.6 ± 1.3 %) compared to the control group (94.4 ± 0.5 %). The biodegradation of ingested PLA was confirmed through the identification of chemical changes in the fragments found in the gut of mealworms, including the formation of oxidized intermediates, alterations in the thermal properties of the
residual polymer, and modifications in specific chemical bonds.

These findings suggest that certain biodegradable post-consumer
plastics could, at least partially, replace traditional components of the
mealworm diet. This substitution could enhance insect growth and
survival while contributing to the mitigation of plastic waste accumu-
lation through increased biodegradation rates. Supporting this idea,
Jung et al. [331] demonstrated that PBS residues in the droppings of
*Zophobas atratus* exhibited a 55.6 % reduction in number-average mo-
lecular weight (Mn) compared to the original PBS foam (14.553 kDa ±
1.12 vs. 8.092 kDa ± 0.782), indicating cleavage of PBS macromolec-
ular chains by gut microbes and/or digestive enzymes.

Furthermore, intestinal juice extracted from *Z. atratus* fed with PBS foam was added to PBS samples in a respirometric chamber maintained at 58 °C. The addition of this gut extract significantly enhanced the mineralization rate of PBS foam, reaching 56.4 % after 45 days—an increase of nearly 1.4 times compared to the control group (41.1 %), which did not receive the enzymatic additive. This effect is likely due to the enrichment of the degradation medium with specific PBS-degrading bacterial strains, such as *Klebsiella oxytoca*. However, it remains to be determined whether a similar synergistic effect on PBS mineralization would also occur under mesophilic conditions.

# 10. Environmental aspects

## 10.1. Nano- and microplastic particles

The fragmentation of plastic products results in the release of nano-
and microplastic particles into the environment [333]. These particles
can exert toxic effects on living organisms that meet them, primarily due
to their ability to adsorb organic pollutants from their surroundings—an
effect linked to their specific physicochemical surface properties.
Moreover, such particles can biomagnify through trophic chains,
increasing their impact at higher levels of the food web [334–336].

Under optimal conditions, the biodegradation process leads biode-
gradable plastics to complete breakdown into carbon dioxide (CO₂), H₂O
and microbial biomass [337]. However, as previously discussed, the
effectiveness of this process varies under real decentralized composting
conditions and strongly depends on environmental factors (e.g. tem-
perature, humidity, aeration, C/N ratio, microbial diversity), which are
often suboptimal in such settings. As a result, biodegradable plastic
products certified as compostable under laboratory conditions (e.g. NF
T51-800) may fail to fully biodegrade under real home-composting
conditions, potentially leading to the release of residual nano- and
microplastics into the decomposition environment [39]. Wei et al. [338]
even observed that biodegradable plastics have a higher potential for
nano- and microplastic generation than conventional plastics, due to
more pronounced fragmentation during their degradation process in
aquatic environments. Thus, the accumulation of nano- and micro-
plastics derived from biodegradable plastics during home composting is
raising increasing concerns, particularly regarding their long-term
impact on soil health when such compost is applied as an organic fer-
tilizer [339].

Indeed, these residues can then be transferred to the soil during
compost application, where they may be ingested by soil fauna [340] or
adsorb organic and inorganic contaminants [341], thereby disrupting
soil biological functions, and/or affect plant physiology.

Over the past decade, increasing attention has been devoted to
evaluating the toxicity of microparticles derived from biodegradable
plastics on living organisms, and to comparing their effects with those of
conventional microplastics [342]. However, contrasting results have
emerged across studies, largely due to the lack of standardized methods
for assessing microplastic toxicity. Critical variables—such as particle
size and shape, concentration, and the diversity of polymer types tes-
ted—often vary considerably between studies, thereby complicating
comparisons and leading to inconsistent findings.

In relation to animal trials, some authors have reported that

biodegradable microplastics can exert toxic effects—such as oxidative
stress, reduced fertility, mucosal vacuolization, muscle disorders, and
downregulation of the expression of certain developmental genes—that
are often comparable to those induced by conventional microplastics
[343–347]. In contrast, other studies have not observed such adverse
effects under similar experimental conditions [348–350].

Regarding toxicity tests conducted on plant organisms, a relatively
larger proportion of studies report toxic effects of biodegradable
microplastics on plants compared to animals [342,351–354]. However,
the concentration of microplastics appears to play a critical role in
determining the direction of the response. Indeed, several studies have
observed increases in plant growth-related traits when plants are grown
in environments containing low concentrations of microplastics [355].
For example, Yang et al. [355] reported a significant increase in the dry
weight of maize (Zea mays L.) grown on substrates containing 0.1 % or 1
% (w/w) PLA microplastics, while a notable decrease was observed at a
concentration of 10 % (w/w) compared to the control without added
microplastics.

Conversely, Wang et al. [356] observed greater height growth in
groundnut plants (Arachis hypogaea) at a PLA microplastic concentration
of 10 % compared to 1 %. Interestingly, the 1 % concentration was itself
associated with increased shoot growth relative to the control. These
contrasting findings suggest that plant responses—whether beneficial or
detrimental—to a given type of microplastic may depend not only on the
concentration of particles but also on interspecific genetic differences.

On the other hand, adverse effects on soil functioning have been
reported at concentrations as low as a few hundred particles per kilo-
gram of soil, particularly in soils repeatedly amended with compost or
sludge contaminated with nano- and microplastics [357,358].

These effects include reductions in microbial biomass, inhibition of
key enzymatic activities (e.g. β-glucosidase, urease, dehydrogenase)
[359,360], and disruptions to biogeochemical cycles through oxidative
stress or altered interactions with organic matter. Microplastics can also
interfere with soil physical properties, such as aggregate stability,
porosity, and water retention, making soils more prone to erosion [361].

Further studies should focus on systematically assessing the con-
centration, particle size, and environmental persistence of biodegrad-
able microplastics in soil when evaluating their potential hazards. In
addition, the same variables—particularly concentration and residence
time—should be considered in parallel when investigating the degra-
dation by-products of biodegradable plastics, such as monomers, olig-
omers, and microbial metabolites. These by-products remain largely
unexplored from a toxicological perspective and warrant thorough
investigation [342].

Moreover, ecotoxicity tests are still significantly underutilized on
compost obtained from trials assessing the domestic compostability of
biodegradable plastic materials. Such tests should be systematically
conducted on a range of organisms from different phyla, and should
include, as far as possible, the quantification of nano- and microparticles
and the characterization of their size distribution within the biodegra-
dation medium.

## 10.2. Environmental persistence of biodegradable microplastics

Numerous challenges remain in applying environmental assessment
tools such as life cycle assessment (LCA) to biodegradable plastics. These
include methodological inconsistencies across studies, as well as
inherent limitations of the tool itself [362], which raise concerns about
the validity of extrapolating results. One notable limitation is the high
sensitivity of LCA outcomes to assumptions regarding end-of-life man-
agement scenarios for plastic products, and the inability of current
models to fully capture the heterogeneous environmental impacts
associated with these scenarios [363].

Biodegradability is an intrinsic property of certain materials which,
in the case of plastics, can substantially limit their accumulation in the
environment. This, in turn, may help mitigate the harmful effects
associated with their long-term persistence in biospheric compartments—particularly aquatic environments, where plastic pollution is especially detrimental.

Consequently, the environmental risk posed by the emission of microparticles during the degradation of plastic materials does not rely solely on the inherent toxicity of a given microplastic type. It also depends on two key factors: (i) the number of particles released, and (ii) their persistence in the environment. The longer microplastics remain in environmental compartments, the higher the likelihood of interactions with living organisms, thereby intensifying ecological exposure and pressure [364].

Conventional plastics require degradation timescales that far exceed the duration of a human lifetime. For instance, Albertsson et al. [365] estimated that polyethylene (PE) powder would take approximately 548 years to fully degrade in the environment. The biodegradability of so-called biodegradable plastics in natural settings remains a subject of ongoing debate, as it is highly dependent on the environmental conditions under which the biodegradation process occurs—and, by extension, on the specific end-of-life pathway followed by these polymeric materials.

Nevertheless, it is reasonable to assume that the environmental persistence of nano- and microparticles resulting from the fragmentation of biodegradable plastics is lower than that of particles derived from conventional plastics. This is primarily due to the lower chemical stability of the bonds in biodegradable polymers, which may render them more susceptible to further degradation under environmental conditions.

Degli-Innocenti et al. [366], building upon the work of Saling et al. [367], proposed a simplified method for evaluating and classifying various polymers based on their potential for environmental harm associated with microplastic emissions following environmental degradation. Their approach relies on a mathematical model designed to estimate a parameter termed microplastic emission potential (MPEP), calculated by multiplying the total number of microplastics generated from the fragmentation of a standard unit of a given polymer by the time required for a single microplastic fragment to fully degrade in the environment.

This preliminary model incorporates two experimentally determined input parameters: (i) the fragmentation rate of a standard polymer unit, and (ii) its biodegradation rate under specific environmental conditions. Using this model, the authors compared a commercial biodegradable formulation—composed of starch, biodegradable polyesters, and a bio-based plasticizer (Mater-Bi EF04P; Novamont)—with polyethylene (PE). Their results showed that the commercial formulation exhibited an MPEP value 1827 times lower than that of PE, a difference primarily attributed to its much faster biodegradation rate relative to polyolefins. In other words, a PE object released into the environment would exhibit a microplastic-related environmental harm potential 1827 times greater than a similar object made from the Mater-Bi-based formulation.

While the model includes several inherent limitations—such as the assumption of uniform microplastic size, constant fragmentation and biodegradation rates, and the exclusion of polymer-specific toxicity differences—it nonetheless offers a valuable framework. This approach underscores that, despite these simplifications, microparticles resulting from the fragmentation of biodegradable plastics **should not** be equated with those originating from conventional, non-biodegradable plastics.

# 11. Conclusions and future perspectives

Home-composting represents a decentralized, cost-effective, and environmentally friendly approach to managing biodegradable plastic waste. However, its efficacy is limited by the inherently low biodegradation rates under mesophilic conditions. This review has underscored the critical role of three interdependent factors influencing the home compostability of biodegradable plastics: (i) the intrinsic properties of the material, including its chemical structure, crystallinity, molecular

weight, and the presence of fillers or additives; (ii) the configuration of the composting system, such as volume, aeration, and temperature dynamics; and (iii) the specific characteristics of the biodegradation medium, particularly microbial diversity and nutrient bioavailability.

Among biodegradable polymers, PHAs and starch-based blends demonstrate the highest potential for successful home composting, whereas others—such as PLA and PBAT—require complementary strategies (e.g. pre-treatments, the use of functional additives, or microbial inoculation) to achieve satisfactory mineralization rates. This review also highlights the considerable heterogeneity in experimental protocols across studies, which complicates cross-comparison of results and underscores the need for harmonized, practice-oriented methodologies that reflect real-world household composting conditions. For example, in real-world trials, contextual data are often missing—such as compost pile temperature profiles, geographic location, exposure to weather, and season—further hindering reproducibility. There is a need for better standardization of protocols and reporting practices in studies.

Future research should focus on several key areas. (i) One priority is the design of materials, including biocomposites, with polymer structures optimized to promote hydrolysis and/or oxidation of polymer chains, to reduce their molecular weight below the threshold required for microbial assimilation. (ii) Another important direction is the evaluation of synergistic strategies that combine bioaugmentation, bio-stimulation, and the use of catalytic additives. (iii) Research should also aim to improve small-scale composting systems to promote heat generation and retention, while ensuring effective aeration control. (iv) Evaluation protocols should emphasize actual biodegradation under real home-composting conditions, explicitly considering the impact of potential seasonal fluctuations. This could involve, for example, adapting the specification requirements of existing standards to better reflect climatic and geographic variability. (v) Furthermore, future studies should quantify the fraction of plastic material effectively converted into residual microbial biomass at the end of compostability trials, in addition to CO₂ evolution. This could be achieved either through indirect approaches in simplified (e.g. vermiculite), activated media incorporating compost extracts, or directly within complex composting environments using isotopic labelling (e.g. ¹³C, ¹⁴C) of the carbon originating from the plastic materials.

Ultimately, promoting the home compostability of bioplastics requires not only technical progress, but also coordinated efforts in citizen education, policy incentives, and methodological alignment to ensure environmental integrity and circularity.

# Funding

This work was supported by the French region of Normandy [RIN50: 22 E01400 - 00123402]; and the Unilasalle Polytechnic Institute [Jean-Baptiste Gagne Foundation].

# CRediT authorship contribution statement

**Jules Bellon:** Writing – original draft, Methodology, Data curation, Visualization, Investigation, Conceptualization, Writing – review & editing, Project administration, Formal analysis, Software, Validation.

**Feriel Bacoup:** Supervision, Writing – review & editing, Funding acquisition, Validation. **Aamir Shehzad:** Validation, Writing – review & editing. **Richard Gattin:** Supervision, Writing – review & editing, Funding acquisition, Validation.

# Declaration of competing interest

The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper.
## Acknowledgements

The authors would like to thank and acknowledge the French region of Normandy and the Unilasalle Polytechnic Institute for financial support of this work. Dr Abdo Bou Sarkis is acknowledged with thanks for his assistance with graphic design and artistry.

## Data availability

No data was used for the research described in the article.

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