Microplastics in Sewage Sludge: Worldwide Presence in Biosolids, Environmental Impact, Identification Methods and Possible Routes of Degradation, Including the Hydrothermal Carbonization Process

Abstract

Biomass-to-biofuel conversion represents a critical component of the global transition to renewable energy. One of the most accessible types of biomass is sewage sludge (SS). This by-product from wastewater treatment plants (WWTPs) contains microplastics (MPs) originating from household, industrial and urban runoff sources. Due to their small size (<5 mm) and persistence, MPs present a challenge when they are removed from sewage systems, where they mainly accumulate (~90%). The presence of MPs in SS poses environmental risks when biosolids are applied as fertilizer in agriculture or incinerated for the purpose of energy production. The key problem is the efficient and reliable identification and reduction of MPs in sewage systems, due to the lack of standardized procedures. The reduction methods for MPs might involve physical, chemical, biological, and hydrothermal approaches, including hydrothermal carbonization (HTC). The HTC of SS produces hydrochar (HC), a solid biofuel, and presents a cutting-edge approach that simultaneously addresses secondary microplastic pollution and renewable biomass-derived energy production. In this article, we review briefly the MPs content in biosolids from different countries, and present HTC as a promising method for their removal from SS. In conclusion, HTC (i) effectively reduces the abundance of MPs in biosolids, (ii) produces an improved solid source of energy, and (iii) contributes to circular SS management.

1. Introduction

The rapid growth in global energy consumption suggests an impending surge in energy demand that cannot be met immediately. Throughout the industrial age, fossil fuels played a vital role in advancing economic development. Coal, oil and natural gas were the primary sources of fuel for generating electricity, powering transportation and driving industrial activity, and they are still the main global energy sources, especially for industry [1]. They are non-renewable and contribute significantly to environmental pollution through greenhouse gas (GHG) emissions during combustion. Global fossil fuel consumption rates continue to rise unsustainably [1], and this is the primary cause of anthropogenic climate change. Depleting fossil fuel resources and concerns about climate change underscore the need for renewable and sustainable fuel options [2]. The group that meets those conditions successfully was labelled as biofuels, which accounted for 6.7% of the fuels used in transport in the European Union (EU) in 2020 [3]. In the quest for sustainable energy sources and the mitigation of climate change, biomass has emerged as a promising alternative to traditional fossil fuels. The utilization of biomass for energy production can help to mitigate GHG emissions such as carbon dioxide (CO₂), methane (CH₄), etc. [4], promote rural development and enhance energy security. This source stands out as one of the most promising due to its abundant availability, renewable nature and versatility in applications [5,6]. Biomass includes agricultural or wood residues, energy crops, animal manure, algae, sewage sludge and other waste materials. Due to its high moisture content and various physical and chemical properties, it requires pretreatment before it is used as a fuel. This can be achieved by processes such as drying, pelletization, briquetting, torrefaction or hydrothermal treatment. These processes lead to the generation of more uniform biofuels with improved fuel properties, as well as the potential to reduce CO₂ emissions. Consequently, they promote the transition towards the circular economy. This energy transition is intended to promote sustainable development, lessen reliance on fossil fuels, and tackle environmental issues such as climate change [7]. The concept of the circular economy is seen as an important approach to manage the world’s CO₂ emissions and support the prevention of further climate change impacts [8,9,10].

A specific type of biomass, sewage sludge (SS), is a by-product from wastewater treatment plants (WWTPs). SS is waste generated during typical treatment processes [11]. Its disposal is a significant global environmental issue. Treated sludge is frequently used as fertilizer in agriculture or disposed of in landfills and/or water bodies. However, the large quantities of sludge produced, as well as the many contaminants involved, make these options environmentally unacceptable, primarily due to the presence of a high content of heavy metals, bacteria, viruses, microplastics and pharmaceuticals. On the other hand, this specific organic waste has many beneficial properties. For example, SS is a renewable solid energy source with a higher heating value (HHV) of about 13.6 MJ kg⁻¹ [12,13]. SS is also a source of carbon (C) and hydrogen (H₂) and has the advantage of being CO₂-neutral, which means that it does not contribute to increased emissions (Figure 1) [8,9]. Moreover, after combustion, the remaining ash is enriched with nutrients, including phosphorous (P) and nitrogen (N₂), which can be used as fertilizer in agriculture and, in addition, the high volume of SS is also reduced. The increasing output of SS in recent years, along with the limitations of current disposal methods, emphasizes the need to find alternative ways to manage this organic material. European legislation prohibits the deposition of SS into landfills or water and imposes limits on the content of heavy metals when SS is introduced into soils. The latest trends in the field of biomass and renewable energy highlight the potential for SS as a “green fuel” and promote the use of its potential. According to Eurostat data, in each year since 2021, more than three megatons of sludge from urban WWTPs were produced in the EU [14]. Even though 20% of SS is incinerated, followed by composting (21%), and 10% is landfilled, approximately 29% is still disposed of in agriculture [15].

However, besides these advantages and possible uses, SS also contains many unwanted or harmful compounds that must be taken into account before further application, especially in agriculture. Wastewater contains a variety of pollutants, which determine the composition of the final sludge. Common pollutants found in wastewater include organic matter, nitrogen- and phosphorous-enriched compounds [16], and pathogenic organisms such as bacteria and viruses [17,18]. The concentrations of these pollutants vary depending on their source, the season, and rainfall intensity. In addition, wastewater and, consequently, SS contain metabolites, pharmaceuticals, polycyclic aromatic hydrocarbons (PAHs) [19,20] and some specific compounds, such as MicroPlastics (MPs) [19,21,22], whose presence in the environment has become alarming in recent years. Despite the initial separation of water from sludge during wastewater treatment, raw sludge retains more than 95% of moisture [23]. Therefore, SS pretreatment is an important stage in wastewater management. This includes various techniques, such as thickening, dewatering, chemical treatment and hydrothermal treatments including hydrolysis, carbonization and liquefaction, which are intended to reduce the volume and weight of the sludge, thus facilitating handling, transport and processing [24]. These processes improve the decomposition of organic compounds, reduce pathogens and stabilize the sludge, making it safer for further management, resulting in a material referred to as “biosolid” [25] that can be used for solid remediation or fertilization. However, these processes can encounter some challenges, such as the complexity of the equipment, high operating costs, significant energy consumption and harmful residues, which necessitate further processing. Regarding hydrothermal treatments, HydroThermal Carbonization (HTC) is a process that might solve some of these problems and could be successfully used for sustainable sludge management. This is a relatively new and promising technology providing highly efficient dewatering, disinfection, and stabilization of SS. Furthermore, it has been reported that the HTC of SS results in a positive reduction in certain micropollutant contents, including microplastics.

Concerning the above-mentioned issues, the aim of this paper is to highlight the presence of microplastics in biosolids around the world, while indicating their environmental impact, identification and degradation methods. Moreover, the paper presents the potential use of the hydrothermal carbonization process as an innovative and promising technology for the effective pretreatment of SS in order to improve its physical and chemical properties and successfully decrease microplastic content.

2. Microplastic—A Modern Contaminant in SS

MPs have been present in our environment for more than 50 years [26] and are widely considered among the greatest environmental threats of the modern world. MPs belong to the Emerging Contaminants group, which includes substances that have recently been identified or recognized as potential threats to human health and the environment. The most commonly used plastics are polyethylene (PE), polypropylene (PP), polyvinyl chloride (PVC), polyethylene terephthalate (PET), polystyrene (PS), polycarbonate (PC), polyurethane (PU) and polyamide (PA) [27]. Consequently, their microparticles are most frequently found in the environment [28,29,30,31]. The large scale of MPs pollution is due to the high and growing demand for and production of plastic materials, which exceeded 400 million tonnes in 2022 [32,33]. Currently, European countries are responsible for 15% of the world’s plastics production. More than a half comes from Asia, including about 32% from China [34].

2.1. Characterization of MPs

Polymer microparticles are generally divided into two main groups. Those produced directly in sizes < 5 mm are referred to as primary MPs, which are used in cosmetics, detergents, pharmaceuticals, paints, coatings, etc. The main group of primary MPs, so-called microbeads, are commonly used in personal care products and cosmetics [35]. Secondary MPs are formed as a result of the breakdown of larger fragments under the influence of physical, chemical and biological factors, and include fragments of plastic bags, bottles, packaging or textiles. Secondary MPs may appear in various shapes, forms and colors (Figure 2). Once released, MPs can be transported over long distances by wind, rivers, tides, rainfall, floods, and sewage disposal systems within both aquatic and terrestrial environments [36,37,38]. Their morphological diversity is a result of the properties of the origin polymer and the influence of external factors, which can cause abrasion and surface damage [39,40].

MPs degradation is a complex process. In general, polymers can degrade according to bio [41,42], photo-oxidative [43,44,45], thermal [46,47,48], ozone [41,49,50], mechanical and chemical [37,41], and catalytic mechanisms [42,51,52]. However, the presence of some biofilms, consisting of bacterial colonies, may provide protection against decomposition [53,54,55].

Figure 2. Images of MPs under optical microscopes: (a) transparent and (b) red fibers [56], (c) triangular, transparent film [56], (d) white foam [36], (e) triangular, blue fragment [57], (f) transparent granule [58].

Figure 2. Images of MPs under optical microscopes: (a) transparent and (b) red fibers [56], (c) triangular, transparent film [56], (d) white foam [36], (e) triangular, blue fragment [57], (f) transparent granule [58].

2.2. Interactions and Health Effects between MPs and Organisms

Although polymers are considered to be biochemically neutral materials [59], many independent scientific studies indicate negative effects of MPs on human, animal and plant organisms. These particles contain chemical additives included in the polymer structure, which may leach uncontrolled at any stage during migration [60,61]. MPs can be absorbed by marine and terrestrial organisms, including humans, via oral, inhalation and dermal routes [57,62,63,64]. Some studies suggest that ingested MPs could cause inflammation, oxidative stress, and cellular damage in the human body [65,66,67]. There are also studies confirming the impact of MP exposure on human respiratory and cardiovascular diseases, and endocrine disorders [54,67]. Aquatic organisms are the most exposed to MPs contamination, as they can easily ingest them directly from water [63,68,69,70,71]. It has been proven that MPs cause oxidative damage to the white blood cells of many species of fish, resulting in growth decrease, inflammation, body length decrease and many more effects [69,72,73]. Soil organisms, such as earthworms, can also ingest MPs, leading to bioaccumulation and biomagnification in food chains [39,74,75]. Studies have shown that they can adhere to plant surfaces, penetrate their tissues and accumulate in root systems [76,77,78,79]. Additionally, MPs can behave in a similar manner to vector-like particles for various micropollutants, such as polychlorinated biphenyls and pesticides, as well as many bacteria, fungi and pathogens [64,80,81]. MPs in landfilled waste can migrate into the soil and enter the edible parts of plants, causing reductions in biomass, root length and shoot height [78]. While plants can often metabolize many organic pollutants, nanomaterials such as MPs are frequently stored within plants, instead of being degraded [82]. This may lead to MPs’ accumulation in roots, stems, leaves and fruits and, consequently, their accumulation in the upstream links of the plant trophic chain [83].

Over the past few years, the presence of MPs in the environment has become an area of intensive study for many scientists in the world, who point to the growing threat posed by their prevalence. The threat continues to increase because many MPs (mainly PP-MPs, PE-MPs, PET-MPs and PS-MPs) have been detected in various foods [52,72,76] and drinking water [66,84,85] around the world.

2.3. Legal Status in EU

The rising production of plastic materials, with disproportionate recycling, continuously increases global MPs pollution. It is estimated that without appropriate global policies in this area, the rate of plastic waste entering the environment will reach 44 million tonnes per year by 2060, which is double the rate recorded in 2019 [86]. Therefore, the rational management of plastic recycling is one of the most important challenges in the current global economy. Emerging reports of MPs contamination have prompted a response in the form of new regulations in this regard. Several countries, including the Netherlands and the US, subsequently banned microbeads in cosmetic products in 2014. Soon after this, similar restrictions were introduced jointly by Sweden, Austria, Luxembourg and Belgium, calling for an MPs ban in all EU countries. In 2015, the Microbead-Free Waters Act was implemented by the US and commenced in 2017. Furthermore, the sales of such products were prohibited in 2018. These actions led other countries, such as the UK, South Korea, Australia, and New Zealand, to follow suit. The EU has also recognized MPs pollution as a pervasive environmental issue, requiring regulatory interventions. In 2023, the REACH EU restriction on MPs intentionally added to products was adopted, and a proposal for a regulation on preventing plastic pellet losses in the environment was released. The EU target for a 30% reduction in MPs release by 2023 is in line with the Zero Pollution Action Plan. Accordingly, EU countries implemented the following regulations concerning plastic usage: Circular Economy Action Plan (updated in 2022), Plastic Recycling Regulation 2022/1616, Single-Use Plastics Directive (Directive (EU) 2019/904), “Plastics Strategy” (2018), Waste Framework Directive (Directive 2008/98/EC), and Packaging Waste Directive (Directive 94/62/EC) [87]. Despite many existing and proposed regulations, the recycling rate for plastics is still generally low. In 2020, in the EU, it was approximately 35% [33] and in the US, the country that generates the most plastic waste (more than 50 million tonnes in 2022), recycled less than 10% [88]. The low reuse of plastic is a consequence of many factors, such as the inadequate infrastructure for recycling, limited consumer awareness of and participation in recycling programs, a lack of incentives for businesses to use recycled materials and the prevalence of single-use plastic products and packaging in the market. For these reasons, plastics are often incorrectly processed or stored unattended, contributing to plastic and, consequently, MPs pollution in the environment [33].

2.4. Sources of MPs in the Environment

As mentioned above, sources of MPs are anthropogenic and related mainly to industrial activities and inappropriate management of plastic waste, including e-waste [89]. Figure 3 represents the general fate of MPs in marine and terrestrial environments. The highest concentrations of MPs are detected in aquatic and soil environments [25,27]. This is because most MPs, such as those derived from the laundering of clothes containing plastic fibers, end up in wastewater [74]. Urban and agricultural runoffs, tire wear and road use also contribute to contamination [75,90,91], while wastewater containing MPs is a significant route for their transport [58]. Wastewater and SS from WWTPs are two of the most important sources of MPs pollution in the marine environment [36,64]. The MPs entering the WWTP within the sewage stream are mainly synthetic fibers (80%), microbeads (17%), films (2%) and foams (>1%) [30]. It is estimated that about 90–95% of the MPs within a sewage stream processed through a traditional mechanical–biological WWTP are deposited in SS [92]. Hence, SS is an essential source of MPs, as it can be returned to the terrestrial environment via fertilizers and soil amendments [92,93,94,95,96]. In addition, MPs, to a small extent, are also being released into the marine environment with the outflow from WWTPs to water bodies [36]. During treatment processes, MPs can absorb pollutants and toxic chemicals, making them more harmful to the natural ecosystem [97].

3. SS Contamination by MPs

3.1. Land and Agricultural Use

The biosolid pathway of secondary MPs’ environmental contamination is gaining increasing attention. Recent studies have revealed the ubiquitous presence of MPs, not only in marine and freshwater environments [58,64,98], but also in terrestrial ecosystems, due to the application of biosolids on land [99,100,101]. This is a result of MPs crossing from SS into the soil, which can occur randomly and without visible signs. Land applications create possibilities for the migration of MPs deep into the soil and groundwater [102,103,104]. This phenomenon poses significant concerns for soil and plant health, water quality and ecosystems’ integrity [105]. Worek et al. (2023) indicated that the use of biosolids as fertilizers will contribute to the emission of a significant amount of MPs into the environment [101]. Therefore, the global significance of studying MPs in biosolids addresses critical environmental and public health concerns, as they can pose a threat to ecosystems. In recent years, many experimental studies related to this issue have been performed in different countries. Accordingly, the results of studies of MPS contents in selected SS have been identified across the world. These include developed and BRICS countries (according to European Parliament data from January 2024, the BRICS group includes Brazil, Russia, India, China, South Africa, Egypt, Ethiopia, Iran and the United Arab Emirates), and the results are summarized in

Table 1. Reported MPs content in biosolids from various studies.

Study Number	Country	Size (If Stated), µm	Type of SS	Av. Concentration, Particles g−1 dw	Identification Method	Reference
1	China	>200 <200	Dewatered sludge	2.533 5.16	stereomicroscopy, FTIR	[106]
2	Norway	>50	Stabilized and dewatered sludge Dewatered sludge	19.898 8.237 2.475 2.78 7.966 1.695	nda	[107]
3	Sweden	>300	Dewatered sludge	16.7	FTIR	[108]
4	China	>37	Dewatered sludge	13.787 15.08 37.463	stereomicroscopy, FTIR	[109]
5	Chile	>8	Dewatered and dried sludge	34	stereomicroscopy	[110]
6	Finland	>20	Dewatered and dried sludge Dewatered, dried, anaerobically digested, and stabilized sludge	186.7 76.3	stereomicroscopy, FTIR	[111]
7	Morocco	<500 >2000	Dewatered sludge	36	stereo- and Raman microscopy, Py-GC/MS + staining	[112]
8	China	>37	Dewatered sludge	1.57–56.4 22.7	FTIR	[113]
9	Canada	>1.6	Final biosolid   Digested sludge	14.1   11.469 14.407	stereomicroscopy, FTIR	[114]
10	USA	20–400	Final biosolid	1	optical microscopy, FTIR	[115]
11	Denmark	20–500	Aerobically digested sludge	169	FTIR	[116]
12	Finland	<250 >500	Dewatered sludge Digested sludge	23 170.9	optical microscopy, FTIR	[117]
13	Korea	<106 <306	Dewatered sludge	14.895 9.475 13.2	digital microscopy, FTIR	[118]
14	Sweden	>10	Digested sludge	14.13	stereomicroscopy, FTIR	[119]
15	China	8–1000	Dewatered sludge	220	stereomicroscopy, FTIR	[120]
16	China	80–1700	Dewatered sludge	2.92	optical microscopy	[121]
17	Australia	>25	Digested sludge	52.1	stereomicroscopy, FTIR	[122]
18	China	>1	Dewatered sludge	1.02	metallurgic microscopy, FTIR	[123]
19	Australia	<5000	Final biosolid	75.3 a	Py-GC/MS	[124]
20	China	-	Anaerobically digested sludge	7.5	FTIR, SEM	[125]
21	Turkey	<2000	Final biosolid	32	stereo- and Ramanmicroscopy	[126]
22	Spain	>50	Final biosolid	7	stereomicroscopy, FTIR	[90]
23	England	>50	Anaerobically digested sludge Dewatered sludge	180.7 97.2 74.7	stereomicroscopy, FTIR	[127]
24	China	-	Dehydrated sludge	14 22.36 21.25 29.66 13.06	optical microscopy, FTIR	[128]
25	Denmark	>10	Dewatered sludge *	0.810	stereomicroscopy, FTIR	[129]
26	India	21–294	Dewatered sludge	0.830	stereo-and SEM microscopy, FTIR	[130]
27	Italy	>300	Anaerobically digested sludge	4.7	stereomicroscopy, FTIR	[131]
28	China	>68 ≥900	Dewatered sludge	12.73	stereomicroscopy, FTIR	[132]
29	China	-	Dehydrated dewatered sludge	6.91 2.19 0.23	digital microscopy, FTIR	[133]
30	Finland	>20	Dewatered and anaerobically digested sludge Anaerobically digested sludge	9.379 102 c	stereo- and Raman microscopy	[134]
31	England	25–178	Digested sludge	7.652 0.5 26 2.0628	FTIR	[135]
32	Sweden	>100	Digested sludge	6.36	stereomicroscopy, FTIR	[136]
33	Iran	>37	Digested sludge Dewatered sludge	238 129	stereo- and optical microscopy, Raman spectroscopy	[137]
34	China	>20	Dewatered sludge	240± 31	optical microscopy, Raman spectroscopy	[35]
35	Germany	≥10	Final biosolid	1–24	nda	[138]
36	Ireland	250–4000	Final biosolid	4.196–15.385	stereo- and SEM microscopy, FTIR	[139]
37	Australia	<1000	Digested sludge	0.996	stereo- and SEM microscopy, FTIR	[140]
38	Canada	>1	Digested sludge Digested and dewatered sludge	14.9 4.4	stereomicroscopy, FTIR	[141]
39	Chile	>8	Digested and dewatered sludge	18–41	stereomicroscopy	[110]
40	England	-	Dewatered sludge	2	stereomicroscopy, FTIR	[142]
41	Germany	>10	Dewatered sludge	1–24	optical microscopy, FTIR	[143]
42	China	>25	Dewatered sludge	1.6 0.7	stereomicroscopy, FTIR	[144]
43	Italy	>10	Dewatered and dry sludge	113	stereomicroscopy, FTIR	[145]
44	Netherlands	>300	Dewatered sludge	0.37 0.95 b 0.51 b 0.76 b	nda	[146]
45	Poland	>0.109	Dewatered sludge Digested sludge	15 28 21 6.7 51 62.6	nda	[147]
46	Spain	>25 36–4720 29–2220	Digested sludge Dried and dewatered sludge	165 101 113	stereomicroscopy, FTIR	[148]
47	US	-	Dewatered sludge	4 (only fibres)	optical microscopy	[149]
48	Australia	>20	Dewatered sludge	55.4 73.8 62.2	nda	[150]
49	US	>1.2	Dewatered sludge	12.2	optical microscopy, LDIR	[47]
50	China	-	Dewatered sludge	4.04	stereomicroscopy, FTIR	[151]
51	France	≥2000	Dewatered sludge	36	stereo- and Raman microscopy, Py-GC/MS	[112]
52	Germany	<500	Dewatered sludge	1–24	stereo- and optical microscopy, FTIR	[152]
53	China	>0.45	Dewatered sludge Dried sludge	4.8 2.7 5.4 2.2	optical and SEM microscopy, FTIR	[153]

Av.—average, ^a—mg g⁻¹, ^b—particlesg⁻¹ wet SS, ^c—particlesL⁻¹, nda—no data available, * SS after hydrothermal liquefaction.

. Dewatered, thermally dried, digested, stabilized SS and final biosolids were chosen for examination as they are the most commonly used types for land and energy purposes. The aim of the research was to investigate the extent of the contamination of SS by the presence of MPs, which were identified by different methods. In Figure 4, the seven countries with the highest concentrations of MPs detected in SS are depicted.

In brief, 53 studies containing 63 samples from numerous countries around the world were investigated. The sizes of the identified polymer particles were very different, ranging from 0.45 to 5000 µm. It was noted that MPs content was most commonly tested in dewatered SS. The largest amount of MPs in this group was detected in China. According to two independent studies, in dewatered SS, 240 particles g⁻¹ within the size range of >20 µm and 220 particles g⁻¹ within the size range of 8–1000 µm were found. In dried, dewatered sludge originating from Spain, more than 100 particles larger than 29 µm were found. A high level of MPs contamination was also obtained from anaerobically digested SS samples originating in England (180.7 particles g⁻¹, size > 50 µm) and digested, dewatered and dried SS from Finland (170.9 particles g⁻¹, size > 500 µm, and 186.7 particles g⁻¹, size > 20 µm, respectively).

The lowest MPs contents, of less than 1 particle g⁻¹, were detected in dewatered SS samples from the Netherlands, China and India, and in digested SS samples from Australia and England. Among EU countries, the highest numbers of MPs were detected in samples from Spain, Finland, Denmark and England, with >100 particles g⁻¹. The number of MPs was counted, but the results of the studies presented in Table 1 are widely scattered. Apart from environmental factors, such as the lifestyles of habitants and the presence of industrial facilities, the most important factor is the lack of a standardized method for extracting MPs. Furthermore, researchers took into account different amounts of material for testing and applied different extraction methods. Therefore, the comparison of these results, even within one country, is not fully reliable. However, the undeniable presence of MPs in SS can pose a significant problem for its further processing.

3.2. Energy Production

SS contains approximately 38–60% volatile matter and 22–40% of C, with HHV in the range of 9–18 MJ kg⁻¹ dry mass. This places it below fossil fuels (e.g., about 30%, 65% and 28 MJ kg⁻¹ for coal, respectively) [12]. The incineration of SS is often considered inefficient because it requires additional pre-drying due to its relatively low NCV, which is caused by its high water content. This results in a reduction in the energy yield, since more energy is used for a longer time. However, SS combustion is still one of the main methods of sludge disposal. The presence of MPs in SS, dedicated to energy production, may affect the course of the combustion process. In general, the hydrocarbon chains included within polymer structures serve as an excellent fuel source due to their clean combustion properties [48,154]. Generally, the HHV of plastics exceeds that in SS, reaching above 40 MJ kg⁻¹ [155]; thus, their presence in SS is beneficial in terms of energy production. However, thermal transformations of MPs during sludge combustion might have adverse effects on the environment. During SS incineration, MPs are subjected to high temperatures, which initiates thermal degradation. For example, the first structural changes in PE associated with degradation can occur at around 100 °C [156]. Polymer decomposition during SS combustion may cause the release of some harmful substances, e.g., volatile organic compounds (VOCs), gases such as carbon oxides (CO, CO₂), methane (CH₄), formaldehyde (CH₂O), and particular matter (PM) into the atmosphere [1]. Therefore, while SS combustion remains a prevalent method of sludge disposal and energy recovery, the MPs content can have a significant impact on its safety. Consequently, a potential pre-treatment may substantiate the removal of MPs from SS before combustion to reduce the overall MPs load in SS while improving its fuel properties. However, it is necessary to assess the amount of MPs deposited in the SS. This information can be obtained by several methods.

4. Separation and Identification of Sludge-Based MPs

Prior to identifying MPs from sewage samples, it is necessary to perform several steps, including three main stages: sample collection, separation (extraction) and identification. However, the lack of standardized methods for MPs extraction causes difficulties with the correlation and comparison of results [157]. For solid samples such as SS, soil, or clay, separation protocols are more complicated, due to the high content of mineral and organic substances that must be removed. Generally, for solid, organic samples, the procedure shown in Figure 5 has been developed and frequently used. However, it can be modified to hasten the process, or enhance MPs separation. Firstly, the SS needs to be dried. Next, the organic matter needs to be effectively removed, which can be conducted using chemical solutions, e.g., 30% H₂O₂, Fenton’s reagent, etc. The next step is, typically, density separation using a saturated salt solution (e.g., CaCl₂ or ZnCl₂), where particles are split based on their density differences. Finally, the top layer with the floating MPs is then decanted and filtered through metal or glass fiber (GF) filters measuring <5 mm and transferred to the microscope slide, or imaged under the microscope directly on the filter. The steps should then be repeated until all organic compounds are removed sufficiently.

MPs identification is based on visual observation and the determination of the type of polymer [95]. In general, these methods can be divided into physical and chemical techniques.

4.1. Physical Analysis

Physical analysis is largely based on the determination of the size, abundance, color and shape of MPs. For sufficiently large particles, this can be achieved visually (with the “naked eye”), or by specific equipment. An optical microscope, such as a stereomicroscope (or dissecting microscope), is most commonly used for fast MPs examination. It allows for the visual evaluation of size, shape and color of MPs. Furthermore, it provides a three-dimensional view, with high accuracy and sensitivity, which is particularly useful for observing surface features and the spatial relationships of particles. Light microscopy is also used to observe and measure MPs. To enhance the visibility and distinguishability of MPs under microscopes, fluorescent staining, e.g., by Nile Red, is used in some cases [158]. The pigment attaches to specific MPs particles, distinguishing them from other compounds. However, visual identification techniques have limitations, relating mainly to the subjectivity of the observation and the difficulty in analyzing very small particles (<50 um) [159]. In order to observe such particles, either a confocal microscope or scanning electron microscopy (SEM) is used. Confocal microscopy uses a laser with a specific wavelength to scan the sample and collect emitted light from a single focal plane at a given time. This method significantly reduces out-of-focus light and enhances image sharpness, enabling high-resolution imaging. This means it can produce thin, detailed slices of the sample and create three-dimensional reconstructions from multiple optical sections. The other technique, SEM, allows for the detailed examination of the morphology and texture of MPs. It uses a focused beam of electrons to scan the surface of a sample, providing extreme resolution and detailed, gray-scale images of the surface topography at the nanometer scale. This technique is particularly effective for observing fine surface details. The accuracy of the image can be further improved by coupling it with energy-dispersive spectroscopy (SEM-EDS) [125].

4.2. Chemical Analysis

Although visual analysis is crucial for MPs identification, it may not be sufficient. To examine the compositions and chemical bonds of MPs samples, Fourier-transform infrared spectroscopy (FTIR) and Raman spectroscopy are widely used [160]. FTIR detects the absorption of infrared radiation by each MPs particle (>45 µm size), producing a spectrum that characterizes the molecular absorption and transmission; in this way, each type of polymer can be determined [161]. This technique can be paired with three detection technologies: attenuated total reflection (ATR), focal plane array (FPA) and µ-FTIR, depending on the particle size [139]. For detecting smaller particles (about 1 µm), µ-FTIR and Raman spectroscopy are used. Raman spectroscopy provides detailed molecular information, making it suitable for identifying the polymer types and additives in MPs particles. Despite potential fluorescence interference, it offers high spatial resolution and minimal sample preparation [162]. X-ray photoelectron spectroscopy (XPS) is another analytical, non-destructive technique used to examine the surface and structure of MPs. Under X-rays and high vacuum, the sample’s emitting electrons are collected and data are analyzed to produce spectra of the chemical compounds and bonds of particles. Pyrolysis–gas chromatography–mass spectrometry (Py-GC/MS) is also a developed, highly sensitive technique for MPs analysis. During Py-GC/MS, the sample is heated to a high temperature in an inert atmosphere, and breaks down into volatile compounds. These are then separated in the GC system, based on their chemical properties, and then identified and quantified by an MS detector. However, this method is relatively expensive and samples are irrevocably destroyed during analysis [124,163]. Other methods for identifying MPs have appeared in the literature and include thermogravimetric analysis with differential scanning calorimetry (TGA-DSC), thermal extraction–desorption combined with compressed gas chromatography–mass spectrometry (TED-GC/MS), and laser direct infrared (LDIR) imaging [47,164,165]. They are not used as often due to the less accessible equipment and more complicated procedures and interpretation of data that they involve. Referring to Table 1, MPs were observed under different microscopes and their chemical composition was determined by several techniques. Among the presented studies, the most common methods for identifying MPs and quantifying their number were the stereomicroscope and FTIR. To observe very small particles (<1 µm), both confocal and SEM microscopy with the µ-FTIR technique were applied, which provided relatively good results.

5. Removal of MPs from Sewage Systems

The persistence and durability of MPs makes them difficult to remove from wastewater and SS. In general, a larger size and a lower polymer density ensure better and more efficient MPs removal during traditional wastewater treatment processes [98]. Although WWTPs are not designed to remove such particles, they are eliminated to some extent, along with other solid particles during treatment [22,35]. Many studies have followed the fate of MPs during conventional wastewater treatment processes, but the results vary considerably. This is due to the complex interaction of factors related to the differences in WWTP designs, operational and environmental conditions, human factors and the temporal variability of effluent, as well as sampling and analytical methods and MPs characteristics [133]. The primary treatment employs physical methods, such as screening and sedimentation, to remove larger debris, including some MPs attached to larger fractions. The secondary treatment, through biological processes such as activated sludge and biofilm reactors, further reduces the MPs content by incorporating these particles into biological flocs [115]. Advanced treatment technologies, such as membrane bioreactors, are used to filter out MPs through fine membranes. Additionally, chemical treatments, e.g., coagulation and flocculation, aggregate polymer microparticles into larger, more easily removable clusters. Other, non-typical tertiary treatments, e.g., sand filtration and the dosing of granular activated carbon, provide a final purification step, ensuring minimal MPs presence in the treated effluent. By integrating these processes, the discharge of MPs into the environment by WWTPs can be significantly reduced. Current methods for removing sewage-derived MPs from conventional urban WWTPs can be generally divided into physical, chemical, biological and hydrothermal methods, and they are listed in Figure 6. According to R. Ahmed et al. (2022), physical methods can remove about 40–99% of the MPs from wastewater, with an emphasis on adsorption as the most effective process [159]. The removal rates of chemical processes are very uneven, ranging from 8–98%, although the coagulation–flocculation mechanism provides the best results. Biological processes can generally remove 17–79% of MPs, although biofilters are the most effective methods [159].

Considering the removal of MPs from SS, it has been proven by Xu Z. et al. (2023) that the conventional thermal drying of SS contributes to the fragmentation of MPs and, consequently, increases the environmental risk [153]. Therefore, there is a need to develop technology that not only breaks MPs into smaller particles, but also degrades them effectively. In terms of the effective management of SS, there is a growing interest in hydrothermal reactions, such as thermal hydrolysis pretreatment (THP), hydrothermal carbonization (HTC) and hydrothermal liquefaction (HTL). There are some reports that confirm MPs degradation during hydrothermal SS processing.

6. Hydrothermal Treatments of Sewage Sludge and the Effect on MPs

Innovative hydrothermal reactions can convert organic feedstock (e.g., SS, agricultural residues, food waste) into valuable products that can be used as biofuels. They are usually carried out in sealed steel reactors, and require the application of high temperatures and pressure, sustained above saturation levels, to keep water in its liquid form. These processes directly use moisture from SS as the reaction medium, as it is transformed into free water [166,167]. During this kind of reaction, organic compounds follow various parallel reaction pathways: hydrolysis, dehydration and decarboxylation. In addition, condensation, polymerization, aromatization and specific reactions characteristic of processes such as deamination may occur [167]. In general, hydrothermal reactions effectively remove water from SS, deactivate pathogens and transform it into valuable products. Accordingly, SS, from its original status as waste, becomes a product with many possible applications. As a result, the volume of sludge is reduced, making SS management more efficient, and facilitating the recovery of energy and nutrients. The effectiveness of hydrothermal reactions primarily depends on different factors: the physical and chemical composition of the input material, temperature and residence time, pressure conditions, pH level and others [167]. Hydrothermal treatments of SS mainly include thermal hydrolysis (THP), carbonization (HTC) and liquefaction (HTL). The methods are described below.

6.1. Thermal Hydrolysis Pretreatment (THP)

Thermal hydrolysis is an advanced hydrothermal reaction used for SS processing. This process is typically carried out in steel reactors at temperatures of around 160–180 °C and pressures of 0.6–1 MPa. These conditions cause the breaking down of complex organic matter, increase the solubility of the organic matter and, consequently, ensure that it is more prone to subsequent anaerobic digestion. The final product, digestate, can be used as a soil conditioner or compost, providing valuable nutrients to agricultural land. In terms of the removal of MPs from SS by THP, there are limited data concerning this issue, because THP temperatures are not sufficiently high to significantly degrade polymers, and higher temperatures are required. However, there is one study by Wang Y. et al. (2024), who investigated the reduction in MPs during the combined processes of THP and anaerobic digestion of SS. The results indicated that the MPs content was reduced by 35% [168].

6.2. Hydrothermal Carbonization (HTC)

HTC is a hydrothermal process that runs at elevated temperatures and pressures, in the range of 180–350 °C and <10 MPa, respectively. The reaction time is typically between one and several hours, which leads to the formation of a solid product, hydrochar (HC), suspended in post-processing liquid, and gases (mainly CO₂, but also CH₄ and H₂). The liquid phase is enriched with nitrogen (N₂) and phosphorous (P), which can be recovered [167,169]. However, due to the high content of organic compounds, it may be toxic, reaching up to 45 gO₂L⁻¹ of chemical oxygen demand (COD) for the SS used as feedstock; thus, purification is required in order to advance any further [167]. HC is brittle and hydrophobic, and it contains reduced levels of moisture and volatile matter, as well as an increased ash content, compared to feedstock. Consequently, it has various potential applications. It can be used as a solid fuel for energy generation, a soil amendment to improve fertility, a precursor for the production of activated carbon and carbon-based materials or, potentially, as a sorbent [170]. According to Czerwińska et al. (2024), the volatile matter content in HC was lower by 14.8% and the ash content was increased by 14.2% compared with the corresponding SS [171]. Therefore, the effect of improving fuel properties is evident. In addition, there are several studies indicating a reduction in heavy metals [172,173,174] or PAHs [19,174,175,176,177] from SS used as feedstock. Consequently, the benefits of processing SS using HTC technology include enhanced energy yield, lower energy input from the reduced moisture content and versatile energy products, including post-processing liquid and gas, which maximize the overall energy recovery. Moreover, the process temperatures enables the degradation of MPs and their many toxic pollutants, preventing their possible release into the environment, in case of any subsequent use of the SS. Recently, a few studies have confirmed the removal of MPs during the HTC of SS, with high temperatures indicated as a key degradation factor. This induces the breaking of chemical bonds and cross-linking (e.g., in MP-PP) or hydrolysis and the splitting of ester groups (e.g., in MP-PET) [92]. After HTC, the MPs content was reduced from 56 up to 100%.

6.3. Hydrothermal Liquefaction (HTL)

Another hydrothermal method, which applies the highest temperatures of the three methods described, is HTL. In HTL, the process temperatures are between 280 °C and 400 °C, and the pressures are from 7 to 30 MPa. The reaction time is typically 10 to 60 min, often when using alkaline catalysts, and, in some cases, with a reduction in gases, e.g., CO and H₂ [92]. HTL converts organic feedstock into liquid bio-crude (or bio-oil), which can be further refined into renewable fuels and other by-products, such as a nutrient-rich water phase and solid residues. The bio-oil typically has a HHV range between 30–36 MJ kg⁻¹ and contains 10 to 20% O₂. In contrast, the SS used for HTL usually has a HHV ranging from 10–20 MJ kg⁻¹ and O₂ contents of 30–50% [12,178]. This highlights that improvements in fuel properties occurred. Regarding MPs’ removal during HTL, there is a study by Chand R. et al. (2022), who performed HTL under supercritical conditions (400 °C and 30 MPa), achieving a 76% reduction in the amount of overall MPs and 97% in terms of MPs mass [129]. However, this high removal rate required much more energy than HTC.

6.4. Comparison of Hydrothermal Methods Applied to Pretreat SS in Terms of MPs Reduction

All three methods, THP, HTC, and HTL, present innovative technology for SS processing and management. The characteristics of the methods are presented in

Table 2. The main characteristics of THP, HTC and HTL methods regarding SS.

Aspect/Technology	THP	HTC	HTL
Process Conditions	160–180 °C, 0.6–1 MPa	180–350 °C, <10 MPa	280–400 °C, 7–30 MPa
Main feature	effective reduction and disinfection of SS
Main product	biogas	solid hydrochar	liquid bio-oil (bio-crude)
Requirements of SS pretreatment	mixing and thickening	dewatering, thickening, digestion, or stabilization
Nutrient recovery	recovery of P, N, K from solid phase	recovery of P, N, K from aqueous phase
Energy and moisture content	increase in CH4 yield in subsequent anaerobic digestion	hydrochar with lower energy content compared to HTL bio-crude, characterized by significant moisture content, may need additional drying for fuel purposes, but 5 times less energy required for drying in comparison to SS	upgrading and moisture reduction in bio-crude to meet fuel properties
Operational costs	lower than for HTC	lower than for HTL	the highest energy input due to the highest temperature and pressure applications
Complexity	high-pressure reactors and control systems		high-pressure reactors and control systems with more developed installation for bio-crude refining
Impact on MPs	minor reduction	significant reduction

. It is important to note that the objective of hydrothermal reactions is the disposal and reuse of SS, but the simultaneous removal of MPs is undoubtedly a major benefit, which is applicable in to new environmental pollutants.

In Figure 7, it is depicted that temperature is the main factor of MPs removal. As can be seen, when a higher temperature is applied, a greater removal of MPs is achieved. Due to the fact that HTC runs in mild conditions and is less energy-consuming compared to other SS hydrothermal pretreatment methods, this technology represents the optimal conditions for MPs reduction.

7. The Impact of HTC on Sewage-Derived MPs

As mentioned previously, the HTC of SS generally reduces the MPs content in the solid product. A general scheme of this process is shown in Figure 8. There are three main studies on this effect, which are comprehensively described below.

7.1. MPs Quantity Reduction

According to Jiang C. et al. (2022), HTC caused the degradation of 11 MPs (>1 μm) in SS [92]. HTC reactions were conducted at 170 °C and 220 °C for 0.5 h residence time. A combination of FTIR and Raman methods employed after the extraction of MPs from HC showed that PC, PMMA, PS, PU and PVDF were removed completely at 170 °C, whereas PA, PE, PET, PO, PP and PVC were removed at a rate of 50%. At 220 °C, the general MPs removal rate obtained was about 87%. Therefore, the MPs removal was strongly related to the process temperature. In other words, the higher the temperature, the better the general removal rate was. Longer reaction times allowed for more extensive degradation, while optimal pressure and moisture levels contributed to the general efficiency of the process. Further research, performed by Xu Z. and Bai X. (2022) in China, was focused on SS containing MPs (>30 μm), subjected to HTC at 180 °C, 220 °C and 260 °C with a reaction time of 3 h [179]. Based on the FTIR technique and optical microscope, which were used for identification, it was found that PU and PS were completely removed at 180 °C, and PET and PA at 220 °C. At 260 °C, only PP and PE were still present, but their contents were reduced by about 80% and 56%, respectively, compared to the initial amount. The general MPs removal rate at 260 °C was 79%. However, polymer decomposition during HTC may result in increased toxicity of the SS, due to the breakdown of molecules into hydrocarbon chains. For example, during pyrolysis, PP-MPs are converted into VOCs, e.g., alkanes, alkenes and dienes [180]. Polymer chain disruption at high temperatures can be seen as the dominant mechanism of PP degradation [154]. In this study, the HTC reaction did not increase the toxicity of the SS in terms of hydrocarbon content. Instead, the PP decomposition resulted in a 10-fold lower alkane content in the liquid product. Toxic characterization was also examined by Jun C. et al. (2023). The researchers attempted to determine the toxicity of dissolved organic matter (DOM) resulting from the decomposition of PA, PE and PP (1 mm and 1 g L⁻¹ concentration during thermal treatment of SS) [181]. HTC reactions were conducted at 180 °C and 220 °C for 1 h, respectively. The degradation of the MPs was confirmed by the MP-DOM’s increase in the liquid product compared to its content in water before HTC. This study has also shown the high toxicity of PO-MP, which may increase the general toxicity of SS during hydrothermal treatment.

The described research has shown that, generally, the HTC process can effectively remove MPs from SS. During the HTC of SS containing plastic particles, bond cleavage of polymer particles occurs [182]. In addition, the resulting HC retains carbon skeletons of the original polymer inside, with some residues of these aromatic and aliphatic carbon compounds [183]. The severity of the HTC reaction shows statistical significance for H₂, O₂, N₂, volatile matter and ash content in the HC. Additionally, the duration of the HTC and the temperature during the process were statistically significant factors affecting the sulfur and fixed carbon contents in the HC, respectively. According to Iñiguez M. et al. (2019), HTC promotes the rearrangement and stabilization of C atoms in plastic particles, leading to the enrichment of the HC with additional C, which improves its fuel properties [183]. Furthermore, due to the breakdown of plastic particles, the structure of the HC gains a new, oxygen-containing functional group, e.g., hydroxyl (-OH) and carbonyl (>C=O). These factors can significantly affect the reactivity, properties and potential applications of HC, making it a suitable material for various industrial, energy and environmental applications [184]. However, the specific chemical composition and final structure of HC obtained from SS contaminated with MPs can vary significantly based on several factors. These include the type and amount of MPs present in the sludge, the process temperature applied during the HTC, the pressure maintained and the reaction time [182]. However, there is still not sufficient information to determine the exact mechanism of these transformations.

7.2. Detection of MPs in HC

During HTC reactions, the plastic particles in the SS are degraded and fragmented, and they undergo transformations leading to dissolution in the aqueous phase, thus reducing the quantity in the solid product. Assessment of the degradation of MPs is possible by observing the surface and morphology, determining the functional groups present, and possible changes in the chemical composition of the particle. To be able to evaluate changes in quantity and quality, MPs need to be separated, which is the same as for SS. In order to observe changes in the surface or structure of MPs precisely, and to assess the overall degree of particulate removal during HTC, sensitive identification techniques must be used. For example, µ-FTIR enables the detection of very small MPs that are formed by the breakdown of larger particles.

Table 3. Methods used to identify MPs subjected to hydrothermal treatments.

MPs/Identification Method	SS-Derived MPs	HC-Derived MPs	Reference
PC, PMMA, PS, PU, PVDF, PA, PE, PET, PO, PP, PVC	Identification of MPs > 50 μm: µ-FTIR (Spotlight 200i with Spectrum Two System, PerkinElmer, Inc., Waltham, MA, USA), eight scans in the reflectance mode. Spectra wavelength range: 600–4000 cm−1; magnification: 50×; resolution: 4 cm−1. Results compared with commercial FTIR library in Spectrum IR software (PerkinElmer, Inc., Waltham, MA, USA).   Identification of MPs < 50 μm: confocal Raman spectroscopy (inVia, Renishaw, PLC, Wotton-under-Edge, UK). Wavelength of red laser: 785 nm; grating 10 × 10 μm; magnification: 50×; spectra wavelength range: 100–3200 cm−1; 10% of laser power with 0.5 s exposure. Results baseline-corrected and compared with a standard Raman library (WiRE 5.3, Renishaw, PLC).	Examination of morphology and comparative purposes: SEM (Ultra55, Zeiss, Jena, Germany), Secondary Electron mode. Resolution: 2 nm @1.0 kV. MPs coated in gold.   Examination of functional groups: µ-FTIR (Spotlight 200i with Spectrum Two System, PerkinElmer, Inc., Waltham, MA, USA). Grating: 35 × 35 μm; 32 scans; spectra wavelength range: 600–4000 cm−1; magnification: 50×; resolution: 4 cm−1.   Verification of surface elements and primary element component analysis: XPS (AXIS Ultra DLD, Kratos Analytical Ltd., Tokyo, Japan), Al Kα X-ray source (1486.7 eV). Resolution: 100 eV and 1 eV, and 50.0 eV and 0.1 eV); spectra fitted using Advantage 5.9 (Thermo Fisher Scientific, Waltham, MA USA), binding energies calibration using containment carbon: C1s ¼ 284.50 eV.	Jiang C. et al. (2022) [92]
PET, PA, PP, PE, PS, PU	Identification of MPs > 30 μm: optical microscope (Leica DM2500 and DM2500 LED, Leica Corporation, Wetzlar, Germany). Camera: 5 MPSHD (Leica MC170 HD, Leica Corporation, Wetzlar, Germany).   Identification of MPs > 30 μm: µ-FTIR (Spotlight 200i, PerkinElmer, Inc., Waltham, MA, USA), 24 scans. Spectra wavelength range: 600–4000 cm−1; resolution: 4 cm−1.	Verification of surface elements: XPS (PHI 5000 VersaProbe, ULVAC-PHI, Kanagawa, Japan). Band energy range: 0–1200 eV.   Examination of functional groups: µ-FTIR (Spotlight 200i, PerkinElmer, Inc., Waltham, MA, USA), 24 scans. Spectra wavelength range: 600–4000 cm−1, resolution: 4 cm−1.	Xu Z. and Bai X. (2022) [179]
PE, PP, PA	Examination of surface morphology: SEM, HITACHI, SU8010.   Examination of chemical compositions changes: XPS (ESCALAB 250Xi, Thermo Fisher). Al Kα X-ray source (1486.6 eV); 12.5 eV voltage; energy step: 0.05 eV; dwelling time: 40–50 ms.		Jun C. et al. (2023) [181]

shows the methods that were used for detecting MPs in SS samples (before HTC) and in corresponding HCs (after HTC) in the previously described studies. The lack of standardization of methodologies for handling MPs is also noticeable here. Researchers have used many available methods to identify MPs. Therefore, there are difficulties in correlating the obtained results. Optical microscopy, Raman and µ-FTIR spectroscopy were commonly used for MPs imaging in SS samples, while SEM, XPS and µ-FTIR were applied to detect and investigate their properties in HC matrixes. The µ-FTIR is a well-known technique and it has enabled the effective determination of MPs before and after the reaction.

8. Conclusions

Microplastics in sewage systems tend to accumulate in sewage sludge. Due to their properties and the difficulty of capturing them directly from wastewater, it seems reasonable to focus on removing these particles from sludge. The conversion of microplastics trapped in sewage sludge or biofuel, through hydrothermal carbonization, presents a new and promising method for decreasing plastic pollution with medium energy input compared to other hydrothermal reactions. This method has positive effects on circular sewage sludge management: (i) it enhances sustainable energy production from biomass and (ii) protects terrestrial and aquatic environments from sludge-derived microplastics. This simple and cost-effective process is competitive in comparison to conventional sewage sludge processing, and it can be further improved and optimized to minimize environmental impact in terms of releasing contaminants. For example, using eco-friendly catalysts and adjusting reaction conditions can significantly improve the efficiency and sustainability of the microplastics hydrothermal degradation process. Although the initial results are encouraging, there are still aspects that require additional investigation. The first issue is the potential for scalability in the hydrothermal carbonization of sewage sludge, which needs to be assessed to determine its feasibility for large-scale or continuous use. The second issue is the lack of a standardized, unified method to detect and image microplastics in sewage systems, especially in sewage sludge. The diversity of the amounts of detected microplastics leads to inconsistent results between studies, making it challenging to compare data and assess the true environmental impact. Since the actual scale of this pollution is unknown, it is difficult to estimate the scale of preventing it.

In summary, the degradation of microplastics during the hydrothermal carbonization of sewage sludge offers the dual benefits of microplastics reduction and energy generation. With continued research and development, hydrothermal carbonization technology has the potential to significantly contribute to environmental sustainability while strengthening energy resilience. However, the standardization of analytical methods in detecting microplastics will lead to the obtaining of consistent and reliable results, which will allow for more effective monitoring, regulation and reduction of microplastic pollution. Therefore, these aspects should be undertaken in future research directions for managing microplastics in sewage sludge.