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Home > Books > Environmental Management in Practice

The Effects of Paper Recycling and its Environmental Impact

Submitted: 24 November 2010 Published: 05 July 2011

DOI: 10.5772/23110

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Environmental Management in Practice

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Author Information

Iveta čabalová *.

• Technical University in Zvolen,Faculty of Wood Sciences and Technology, Slovakia

František Kačík

Anton geffert *, danica kačíková *.

*Address all correspondence to:

1. Introduction

It is well known the paper production (likewise the other brands of industry) has enormous effects on the environment. The using and processing of raw materials has a variety of negative effects on the environment.

At the other hand there are technologies which can moderate the negative impacts on the environment and they also have a positive economical effect. One of these processes is the recycling, which is not only the next use of the wastes. The main benefit of the recycling is a double decrease of the environment loading, known as an environmental impact reducing. From the first view point, the natural resources conserves at side of the manufacturing process inputs, from the second view point, the harmful compounds amount leaking to the environment decreases at side of the manufacturing process outputs.

The paper production from the recycled fibers consumes less energy; conserves the natural resources viz. wood and decreases the environmental pollution. The conflict between economic optimization and environmental protection has received wide attention in recent research programs for waste management system planning. This has also resulted in a set of new waste management goals in reverse logistics system planning. Pati et al. (2008 ) have proposed a mixed integer goal programming (MIGP) model to capture the inter-relationships among the paper recycling network system. Use of this model can bring indirectly benefit to the environment as well as improve the quality of waste paper reaching the recycling unit.

In 2005, the total production of paper in Europe was 99.3 million tonnes which generated 11 million tonnes of waste, representing about 11% in relation to the total paper production. The production of recycled paper, during the same period, was 47.3 million tonnes generating 7.7 million tonnes of solid waste (about 70% of total generated waste in papermaking) which represents 16% of the total production from this raw material ( CEPI 2006 ).

The consumption of recovered paper has been in continuous growth during the past decades. According to the Confederation of European Paper Industries (CEPI), the use of recovered paper was almost even with the use of virgin fiber in 2005. This development has been boosted by technological progress and the good price competitiveness of recycled fiber, but also by environmental awareness – at both the producer and consumer ends – and regulation that has influenced the demand for recovered paper. The European paper industry suffered a very difficult year in 2009 during which the industry encountered more down-time and capacity closures as a result of the weakened global economy. Recovered paper utilisation in Europe decreased in 2009, but exports of recovered paper to countries outside CEPI continued to rise, especially to Asian markets (96.3%). However, recycling rate expressed as “volume of paper recycling/volume of paper consumption” resulted in a record high 72.2% recycling rate after having reached 66.7% the year before ( Fig. 1 ) ( Hujala et al. 2010 ;CEPI 2006; European Declaration on Paper Recycling 2010; Huhtala& Samakovlis 2002 ; CEPI Annual Statistic 2010).

European paper recycling 1995-2009 in million tonnes (European Declaration on Paper Recycling 2006 – 2010, Monitoring Report 2009 (2010) (www.erpa.info)

Recycling is not a new technology. It has become a commercial proposition since Matthias Koops established the Neckinger mill, in 1826, which produced white paper from printed waste paper. However, there were very few investigations into the effect of recycling on sheet properties until late 1960's. From then until the late 1970's, a considerable amount of work was carried out to identify the effects of recycling on pulp properties and the cause of these effects ( Nazhad 2005 ; Nazhad& Paszner 1994 ). In the late 1980's and early 1990's, recycling issues have emerged stronger than before due to the higher cost of landfills in developed countries and an evolution in human awareness. The findings of the early 70's on recycling effects have since been confirmed, although attempts to trace the cause of these effects are still not resolved ( Howard &Bichard 1992 ).

Recycling has been thought to reduce the fibre swelling capability, and thus the flexibility of fibres. The restricted swelling of recycled fibres has been ascribed to hornification, which has been introduced as a main cause of poor quality of recycled paper ( Scallan&Tydeman 1992 ). Since 1950's, fibre flexibility among the papermakers has been recognized as a main source of paper strength. Therefore, it is not surprising to see that, for over half a century, papermakers have supported and rationalized hornification as a main source of tensile loss due to drying, even though it has never been fully understood ( Sutjipto et al. 2008 ).

Recycled paper has been increasingly produced in various grades in the paper industry. However, there are still technical problems including reduction in mechanical strength for recycled paper. Especially, chemical pulp-origin paper, that is, fine paperrequires a certain level of strength. Howard & Bichard (1992 ) reported that beaten bleachedkraft pulp produced handsheets which were bulky and weak in tensile and burst strengthsby handsheet recycling. This behaviour could be explained by the reduction in re-swelling capability or the reduction in flexibility of rewetted pulp fibers due to fiber hornification and, possibly, by fines loss during recycling processes, which decrease both total bondingarea and the strength of paper ( Howard 1995 ; Nazhad&Paszner 1994 ; Nazhad et al. 1995 ; Khantayanuwong et al.2002 ; Kim et al. 2000 ).

Paper recycling is increasingly important for the sustainable development of the paper industry as an environmentally friendly sound. The research related to paper recycling is therefore increasingly crucial for the need of the industry. Even though there are a number of researches ascertained the effect of recycling treatment on properties of softwood pulp fibres ( Cao et al. 1999 ; Horn 1975 ; Howard&Bichard 1992 ; Jang et al. 1995 ), however, it is likely that hardwood pulp fibres have rarely been used in the research operated with recycling treatment. Changes in some morphological properties of hardwood pulp fibres, such as curl, kink, and length of fibre, due to recycling effects also have not been determined considerably. This is possibly because most of the researches were conducted in the countries where softwood pulp fibres are commercial extensively ( Khantayanuwong 2003 ). Therefore, it is the purpose of the present research to crucially determine the effect of recycling treatment on some important properties of softwood pulp fibres.

2. Alterations of pulp fibres properties at recycling

The goal of a recycled paper or board manufacturer is to make a product that meets customers΄ specification and requirements. At the present utilization rate, using recycled fibres in commodity grades such as newsprint and packaging paper and board has not caused noticeable deterioration in product quality and performance ( Čabalová et al. 2009 ). The expected increase in recovery rates of used paper products will require a considerable consumption increase of recycled fibres in higher quality grades such as office paper and magazine paper. To promote expanded use of recovered paper, understanding the fundamental nature of recycled fibres and the differences from virgin fibres is necessary.

Essentially, recycled fibres are contaminated, used fibres. Recycled pulp quality is, therefore, directly affected by the history of the fibres, i.e. by the origins, processes and treatments which these fibres have experienced.

McKinney (1995) classified the history into five periods:

fibre furnish and pulp history

paper making process history

printing and converting history

consumer and collection history

recycling process history.

To identity changes in fibre properties, many recycling studies have occurred at laboratory. Realistically repeating all the stages ofthe recycling chain is difficult especially when including printing and deinking. Some insight into changes in fibre structure, cell wall properties, and bonding ability is possible from investigations using various recycling procedures, testing methods, and furnishes.

Mechanical pulp is chemically and physically different from chemical pulp then recycling effect on those furnishes is also different. When chemical fibres undergo repeated drying and rewetting, they are hornified and can significantly lose their originally high bonding potential ( Somwand et al. 2002 ; Song & Law 2010 ; Kato & Cameron 1999 ; Bouchard & Douek 1994 ; Khantayanuwong et al. 2002 ; Zanuttini et al. 2007 ; da Silva et al. 2007 ). The degree of hornification can be measured by water retention value (WRW) ( Kim et al. 2000 ). In contrast to the chemical pulps, originally weakermechanical pulps do not deteriorate but somewhat even improve bonding potential during a corresponding treatment. Several studies( Maloney et al. 1998 ; Weise 1998 ; Ackerman et al. 2000 ) have shown good recyclability of mechanical fibres.

Adámková a Milichovský (2002 ) present the dependence of beating degree ( SR –Schopper-Riegler degree) and WRV from the relative length of hardwood and softwood pulps. From their results we can see the WRV increase in dependence on the pulp length alteration is more rapid at hardwood pulp, but finally this value is higher at softwood pulps. Kim et al. (2000 ) determined the WRV decrease at softwood pulps with the higher number of recycling (at zero recycling about cca 1.5 g/g at fifth recycling about cca 1.1 g/g).Utilisation of the secondary fibres to furnish at paper production decrease of the initial need of woody raw (less of cutting tress) but the paper quality is not significantly worse.

2.1. Paper recycling

The primary raw material for the paper production is pulps fibres obtaining by a complicated chemical process from natural materials, mainly from wood. This fibres production is very energy demanding and at the manufacturing process there are used many of the chemical matters which are very problematic from view point of the environment protection. The suitable alternative is obtaining of the pulp fibres from already made paper. This process is far less demanding on energy and chemicals utilisation. The paper recycling, simplified, means the repeated defibring, grinding and drying, when there are altered the mechanical properties of the secondary stock, the chemical properties of fibres, the polymerisation degree of pulp polysaccharidic components, mainly of cellulose, their supramolecular structure, the morphological structure of fibres, range and level of interfibres bonds e.g.. The cause of above mentioned alterations is the fibres ageing at the paper recycling and manufacturing, mainly the drying process.

At the repeat use of the secondary fibres, it need deliberate the paper properties alter due to the fiber deterioration during the recycling, when many alteration are irreversible. The alteration depth depends on the cycle’s number and way to the fibres use. The main problem is the decrease of the secondary pulp mechanical properties with the continuing recycling, mainly the paper strength ( Khantayanuwong et al. 2002 ; Jahan 2003 ; Hubbe & Zhang 2005 ; Garg & Singh 2006 ; Geffertová et al. 2008 ; Sutjipto et al. 2008 ). This decrease is an effect of many alterations, which can but need not arise in the secondary pulp during the recycling process. The recycling causes the hornification of the cell walls that result in the decline of some pulp properties. It is due to the irreversible alterations in the cells structure during the drying ( Oksanen et al. 1997 ; Kim et al. 2000 ; Diniz et al. 2004 ).

The worse properties of the recycled fibres in comparison with the primary fibres can be caused by hornification but also by the decrease of the hydrophilic properties of the fibres surface during the drying due to the redistribution or migration of resin and fat acids to the surface ( Nazhad& Paszner 1994 ; Nazhad 2005 ). Okayama (2002 ) observed the enormous increase of the contact angle with water which is related to the fiber inactivation at the recycling. This process is known as „irreversible hornification“.

Paper recycling saves the natural wood raw stock, decreases the operation and capital costs to paper unit, decrease water consumption and last but not least this paper processing gives rise to the environment preservation (e.g. 1 t of waste paper can replace cca 2.5 m 3 of wood).

A key issue in paper recycling is the impact of energy use in manufacturing.Processing waste paper for paper and board manufacture requires energy that isusually derived from fossil fuels, such as oil and coal. In contrast to the productionof virgin fibre-based chemical pulp, waste paper processing does not yield a thermalsurplus and thus thermal energy must be supplied to dry the paper web. If,however, the waste paper was recovered for energy purposes the need for fossil fuelwould be reduced and this reduction would have a favourable impact on the carbondioxide balance and the greenhouse effect. Moreover, pulp production based onvirgin fibres requires consumption of round wood and causes emissions of air-pollutingcompounds as does the collection of waste paper. For better paper utilization, an interactive model, the Optimal Fibre Flow Model, considersboth a quality (age) and an environmental measure of waste paper recycling was developed ( Byström&Lönnstedt 1997 ).

2.1.1. Influence of beating on pulp fibres

Beating of chemical pulp is an essential step in improving the bonding ability of fibres. The knowledge complete about beating improves the present opinion of the fibres alteration at the beating. The main and extraneous influences of the beating device on pulps were defined.The main influences are these, each of them can be improve by the suitable beating mode, but only one alteration cannot be attained. Known are varieties of simultaneous changes in fibres, such as internal fibrilation, external fibrilation, fiber shortening or cutting, and fines formation ( Page 1989 ; Kang & Paulapuro 2006a ; Kang & Paulapuro 2006c ).

Freeing and disintegration of a cell wall affiliated with strongswelling expressed as an internal fibrilation and delamination. The delamination is a coaxial cleavage in the middle layer of the secondary wall.It causes the increased water penetration to the cell wall and the fibre plasticizing.

External fibrillation and fibrils peeling from surface, which particularly or fully attacks primary wall and outside layers of secondary walls.Simultaneously from the outside layers there arecleavage fibrils, microfibrils, nanofibrils to the macromolecule of cellulose and hemicelluloses.

Fibres shortening in any place in any angle-wise across fibre in accordance with loading, most commonly in weak places.

Concurrently the main effects at the beating also the extraneous effects take place, e.g. fines making, compression along the fibres axis, fibres waving due to the compression. It has low bonding ability and it influences the paper porosity,stocks freeness ( Sinke&Westenbroek 2004 ).

The beating causes the fibres shortening, the external and internal fibrillation affiliated with delamination and the fibres plasticizing. The outside primary wall of the pulp fibre leaks water little, it has usually an intact primary layer and a tendency to prevent from the swelling of the secondary layer of the cell wall. At the beating beginning there are disintegrated the fibre outside layers (P and S1), the fibrilar structure of the fibre secondary layer is uncovering, the water approach is improving, the swelling is taking place and the fibrillation process is beginning. The fibrillation process is finished by the weaking and cleavaging of the bonds between the particular fibrils and microfibrils of cell walls during the mechanical effect and the penetration into the interfibrilar spaces, it means to the amorphous region, there is the main portion of hemicelluloses.

Češek& Milichovský (2005 ) showed that with the increase of pulp beating degree the standard rheosettling velocity of pulp decreases more at the fibres fibrillation than at the fibres shortening.

Refining causes a variety of simultaneous changes in the fiber structure, such as internal fibrillation, external fibrillation and fines formation. Among these effects, swelling is commonly recognized as an important factor affecting the strength of recycled paper ( Kang & Paulapuro 2006d ).

Scallan & Tigerstrom (1991 ) observed the elasticity modulus of the long fibres from kraft pulp during the recycling. Flexibility decrease was evident at the beating degree decrease ( SR), and also with the increase of draining velocity of low-yield pulp.

Alteration of the breaking length of the paper sheet drying at the temperature of 80, 100 a 120°C during eightfold recycling

The selected properties of the pulp fibres and the paper sheets during the process of eightfold recycling at three drying temperatures of 80, 100, 120°C.

From the result on Fig. 2 we can see the increase of the pulp fibres active surface takes place during the beating process, which results in the improve of the bonding and the paper strength after the first beating. It causes also the breaking length increase of the laboratory sheets. The secondary fibres wear by repeated beating, what causes the decrease of strength values ( Table 1 ).

The biggest alterations of tear index ( Fig. 3 ) were observed after fifth recycling at the bleached softwood pulp fibres. The first beating causes the fibrillation of the outside layer of the cell wall, it results in the formation of the mechanical (felting) and the chemical bonds between the fibres. The repeated beating and drying dues, except the continuing fibrillation of the layer, the successive fibrils peeling until the peeling of the primary and outside secondary layer of the cell wall. It discovers the next non-fibriled layer S2 (second, the biggest layer of the secondary wall) what can do the tear index decrease. The next beating causes also this layer fibrillation, which leads to the increase of the strength value ( Fig. 3 , Tab. 1 ).Paper strength properties such as tensile strength and Scott bond strength were strongly influenced by internal fibrillation; these could also be increased further by promoting mostly external fibrillation ( Kang & Paulapuro 2006b ).

The course of the breaking length decrease and the tearing strength increase of the paper sheet is in accordance with the results of Sutjipto et al. (2008 ) at the threefold recycling of the bleached (88% ISO) softwood pulps prepared at the laboratory conditions, beated on PFI mill to 25 SR.

Tear index alteration of the paper sheets drying at the temperature of 80, 100 a 120°C, during eightfold recycling

Song & Law (2010 ) observedkraft pulp oxidation and its influence on recycling characteristics of fibres, the found up the fibre oxidation influences negatively the tear index of paper sheets.Oxidation of virgin fibre prior to recycling minimized the loss of WRV and sheet density.

The beating causes the fibres shortening and fines formation which is washed away in the large extent and it endeds in the paper sludges. This waste can be further processed and effective declined.

Within theEuropean Union several already issued and other foreseendirectives have great influence on the waste managementstrategy of paper producing companies. Due to the large quantities ofwaste generated, the high moisture content of the wasteand the changing composition, some recovery methods,for example, conversion to fuel components, are simplytoo expensive and their environmental impact uncertain.The thermal processes, gasification and pyrolysis, seem tobe interesting emerging options, although it is still necessaryto improve the technologies for sludge application.Other applications, such as the hydrolysis to obtain ethanol,have several advantages (use of wet sludge and applicabletechnology to sludges) but these are not welldeveloped for pulp and paper sludges. Therefore, at thismoment, the minimization of waste generation still hasthe highest priority ( Monte et al. 2009 ).

2.1.2. Drying influence on the recycled fibres

Characteristic differences between recycled fibres and virgin fibres can by expected. Many of these can by attributed to drying. Drying is a process that is accompanied by partially irreversible closure of small pores in the fibre wall, as well as increased resistance to swelling during rewetting. Further differences between virgin and recycled fibres can be attributed to the effects of a wide range of contaminating substances ( Hubbe et al. 2007 ). Drying, which has an anisotropic character, has a big influence on the properties of paper produced from the secondary fibres.During the drying the shear stress are formatted in the interfibrilar bonding area. The stresses formatted in the fibres and between them effect the mechanical properties in the drying paper. The additional effect dues the tensioning of the wet pulp stock on the paper machine.

During the drying and recycling the fibres are destructed. It is important to understand the loss of the bonding strength of the drying chemical fibres. Dang (2007 ) characterized the destruction like a percentage reduction of ability of the water retention value (WRV) in pulp at dewatering.

Hornification = [(WRV 0 -WRV 1 )/WRV 0 ]. 100 [%],

WRV 0 –is value of virgin pup

WRV 1 –the value of recycled pulp after drying and reslushing.

According to the prevailing concept, hornification occurs in the cell wall matrix of chemical fibres. During drying, delaminated parts of the fiber wall, i.e., cellulose microfibrils become attached as Fig. 4 shows ( Ackerman et al. 2000 ).

Changes in fiber wall structure ( Weise &Paulapuro 1996 )

Shrinkage of a fiber cross section ( Ackerman et al. 2000 )

Hydrogen bonds between those lamellae also form. Reorientation and better alignment of microfibrils also occur. All this causes an intensely bonded structure. In a subsequent reslushing in water, the fiber cell wall microstructure remains more resistant to delaminating forces because some hydrogen bonds do not reopen. The entire fiber is stiffer and more brittle ( Howard 1991 ). According to some studies ( Bouchard &Douek 1994 ; Maloney et al. 1998 ), hornification does not increase the crystallinity of cellulose or the degree of order in the hemicelluloses ofthe fiber wall.

The drying model of Scallan ( Laivins&Scallan 1993 ) suggests that hornification prevents the dry structure in A from fully expanding to the wet structure in D. Instead, only partial expansion to B may be possible after initial drying creates hydrogen bonds between the microfibrils( Kato & Cameron 1999 )

Weise & Paulapuro (1996 ) did very revealing work about the events during fiber drying. They studied fiber cross section of kraft fibers in various solids by Confocal Laser Scanning Microscope (CLSM) and simultaneously measured hornification with WRV tests. Irreversible hornification of fibers began on the degree of beating. It does not directly follow shrinkage since the greatest shrinkage of fibers occurs above 80 % solids content. In Figs. 4 and 5 , stage A represented wet kraft fiber before drying. In stage B, the drainage has started tocause morphological changes in the fiber wall matrix at about 30 % solids content. The fiber wall lamellae start to approach each other because of capillary forces. During this stage, the lumen can collapse. With additional drying, spaces between lamellae continue shrinking to phase C where most free voids in the lamellar structure of the cell wall have already closed. Toward the end of drying in stage D, the water removal occurs in the fine structure of the fiber wall. Kraft fiber shrink strongly and uniformly during this final phase of drying, i.e., at solid contents above 75-80 %. The shrinkage of stage D is irreversible.

At a repeated use of the dried fibres in paper making industry, the cell walls receive the water again. Then the opposite processes take place than in the Fig. 4 and 5 . It show Scallan´s model of the drying in Fig. 6 .

The drying dues also macroscopic stress applied on paper and distributed in fibres system according a local structure.

2.1.3. Properties of fibres from recycled paper

The basic properties of origin wet fibres change in the drying process of pulp and they are not fully regenerated in the process of slushing and beating.

The same parameters are suitable for the description of the paper properties of secondary fibres and fibres at ageing as well as for description of primary fibres properties. The experiences obtained at the utilisation of waste paper showed the secondary fibres have very different properties from the origin fibres. Next recycling of fibres causes the formation of extreme nonhomogeneous mixture of various old fibres. At the optimum utilisation of the secondary fibres it need take into account their altered properties at the repeated use. With the increase number of use cycles the fibres change irreversible, perish and alter their properties. Slushing and beating causes water absorption, fibres swelling and a partial regeneration of properties of origin fibres. However the repeated beating and drying at the multiple production cycles dues the gradual decrease of swelling ability, what influences a bonding ability of fibres. With the increase of cycles number the fibres are shortened. These alterations express in paper properties. The decrease of bonding ability and mechanical properties bring the improving of some utility properties. Between them there is higher velocity of dewatering and drying, air permeability and blotting properties improve of light scattering, opacity and paper dimensional stability.

The highest alterations of fibres properties are at the first and following three cycles. The size of strength properties depends on fibres type ( Geffertová et al. 2008 ).

Drying influences fibres length, width, shape factor, kinks which are the important factors to the strength of paper made from recycled fibres. The dimensional characteristics are measured by many methods, known is FQA (Fiber Quality Analyser), which is a prototype IFA (Imaging Fiber Analyser) and also Kajaani FS-200 fibre-length analyser. They measure fibres length, different kinks and their angles. Robertson et al. (1999 ) show correlation between methods FQA and Kajaani FS-200. A relatively new method of fibres width measurement is also SEM (Scanning Electron Microscope) ( Bennis et al. 2010 ). Among devices for analyse of fibres different properties and characteristics, e.g. fibres length and width, fines, various deformations of fibres and percentage composition of pulp mixture is L&W Fiber Tester (Lorentzen & Wettre, Sweden). At every measurement the minimum of 20 000 fibres in a sample is evaluated. On Fig. 7 there is expressed the alteration of fibres average length of softwood pulps during the eightfold recycling at the different drying temperature of pulp fibres.

Influence of recycling number and drying temperature on length of softwood pulps

Influence of recycling number and drying temperature on width of softwood pulps

The biggest alteration were observed after first beating (zero recycling), when the fibres average length decrease at the sheet drying temperature of 80°C about 17%, at the temperature of 100°C about 15.6% and at the temperature of 120°C about 14.6%.

After the first beating the fibres average width was markedly increased at the all temperatures dues to the fibrillation influence. The fibres fibrillation causes the fibre surface increase. Following markedly alteration is observed after fifth recycling, when the fibres average width was decreased. We assume the separation of fibrils and microfibrils from the cell walls dues the separation of the cell walls outside layer, the inside nonfibriled wall S2 was discovered and the fibres average width decreased. After the fifth recycling the strength properties became worse, mainly tear index ( Fig. 3 ).

The softwood fibres are longer than hardwood fibres, they are not so straight. The high value of shape factor means fibres straightness. The biggest alterations of shape factor can be observed mainly at the high drying temperatures. The water molecules occurring on fibres surface quick evaporate at the high temperatures and fibre more shrinks. It can result in the formation of weaker bonds between fibres those surfaces are not enough near. At the beginning of wet paper sheet drying the hydrogen bond creates through water layer on the fibres surface, after the drying through monomolecular layer of water, finally the hydrogen bond results after the water removal and the surfaces approach. It results in destruction of paper and fibre at the drying.

Chemical pulp fines are an important component in papermaking furnish. They can significantly affect the mechanical and optical properties of paper and the drainage properties of pulp ( Retulainen et al. 1993 ). Characterizing the fines will therefore allow a better understanding of the role of fines and better control the papermaking process and the properties of paper. Chemical pulp fines retard dewatering of the pulp suspension due to the high water holding capacity of fines. In the conventional method for characterizing the role of fines in dewatering, a proportion of fines is added to the fiber furnish, and then only the drainage time. Fines suspension is composed of heterogeneous fines particles in water. The suspension exhibits different rheological characteristics depending on the degree of interaction between the fines particles and on their hydration ( Kang & Paulapuro 2006b ).

From Fig. 9 we can see the highest formation of fines were after seventh and eight recycling, when the fibres were markedly weakened by the multiple using at the processes of paper making. They are easier and faster beating (the number of revolution decreased by the higher number of the recycling).

Influence of recycling process and drying temperature on pulp fines changes

The macroscopic level (density, volume, porosity, paper thickness) consists from the physical properties very important for the use of paper and paperboard. They indirectly characterize the three dimensional structure of paper ( Niskanen 1998 ). A paper is a complex structure consisting mainly of a fibre network, filler pigment particles and air. Light is reflected at fibre and pigment surfaces in the surface layer and inside the paper structure. The light also penetrates into the cellulose fibres and pigments, and changes directions. Some light is absorbed, but the remainder passes into the air and is reflected and refracted again by new fibres and pigments. After a number of reflections and refractions, a certain proportion of the light reaches the paper surface again and is then reflected at all possible angles from the surface. We do not perceive all the reflections and refractions (the multiple reflections or refractions) which take place inside the paper structure, but we perceive that the paper has a matt white surface i.e. we perceive a diffuse surface reflection. Some of the incident light exists at the back of the paper as transmitted light, and the remainder has been absorbed by the cellulose and the pigments. Besides reflection, refraction and absorption, there is a fourth effect called diffraction. In other contexts, diffraction is usually the same thing as light scattering, but within the field of paper technology, diffraction is only one aspect of the light scattering phenomenon. Diffraction occurs when the light meets particles or pores which are as large as or smaller then the wavelength of the light, i.e. particles which are smaller than one micrometer (μm). These small elements oscillate with the light oscillation and thus function as sites for new light sources. When the particles or pores are smaller than half of the light wavelength the diffraction decreases. It can be said that the light passes around the particle without being affected ( Pauler 2002 ).

The opacity, brightness, colouring and brilliance are important optical properties of papers and paperboards. For example the high value of opacity is need at the printing papers, but opacity of translucent paper must be lower. The paper producer must understand the physical principles of the paper structure and to determine their characteristics composition. It is possible to characterize nondirect the paper structure. The opacity characterizes the paper ability to hide a text or a figure on the opposite side of the paper sheet. The paper brightness is a paper reflection at a blue light use. The blue light is used because the made fibers have yellowish colour and a human eye senses a blue tone like a white colour.The typical brightness of the printing papers is 70 – 95% and opacity is higher than 90% ( Niskanen 1998 ).

3. Paper ageing

The recycled paper is increasingly used not only for the products of short term consumption (newspaper, sanitary paper, packaging materials e.g.), but also on the production of the higher quality papers, which can serve as a culture heritage medium. The study of the recycled papers alterations in the ageing process is therefore important, but the information in literature are missing.

The recycling is also another form of the paper ageing. It causes the paper alterations, which results in the degradation of their physical and mechanical properties. The recycling causes a chemical, thermal, biological and mechanical destruction, or their combination ( Milichovský 1994 ; Geffertová et al. 2008 ).The effect of the paper ageing is the degradation of cellulose, hemicelluloses and lignin macromolecules, the decrease of low molecular fractions, the degree of polymerisation (DP) decrease, but also the decline of the mechanical and optical properties ( El Ashmawy et al. 1974 ; Valtasaari & Saarela 1975 ; Lauriol et al. 1987a ,b,c; Bansa 2002 ; Havermans 2003 ; Dupont & Mortha 2004 ; Kučerová & Halajová, 2009 ; Čabalová et al. 2011 ).Cellulose as the most abundant natural polymer on the Earth is very important as a renewable organic material. The degradation of cellulosebasedpaper is important especially in archives and museums where ageing in various conditions reduces the mechanical properties and deteriorates optical quality of stored papers, books and other artefacts. The low rate of paper degradation results in the necessity of using accelerating ageing tests. The ageing tests consistin increasing the observed changes of paper properties, usually by using different temperature, humidity, oxygen content and acidity, respectively. Ageing tests are used in studies of degradation rate and mechanism. During the first ageing stages—natural or accelerated—there are no significant variations in mechanical properties: degradation evidence is only provided by measuring chemical processes. Oxidation induced by environmental conditions, in fact, causes carbonyl and carboxyl groups formation, with great impact on paper permanence and durability, even if mechanical characteristics are not affected in the short term ( Piantanida et al. 2005 ). During the degradation two main reactions prevail – hydrolysis of glycosidic bonds and oxidation of glucopyranose rings. As a result of some oxidation processes keto- and aldehyde groups are formed. These groups are highly reactive; they are prone to crosslinking, which is the third chemical process of cellulose decay ( Bansa 2002 , Calvini & Gorassini 2006 ).

At the accelerated paper ageing the decrease of DP is very rapid in the first stages of the ageing, later decelerates. During the longer time of the ageing there was determined the cellulose crosslinking by the method of size exclusion chromatography (SEC) ( Kačík et al. 2009 ). The similar dependences were obtained at the photo-induced cellulose degradation ( Malesic et al. 2005 ).

An attention is pay to the kinetic of the cellulose degradation in several decades, this process was studied by Kuhn in 1930 and the first model of the kinetic of the cellulose chains cleavage was elaborated by Ekenstam in 1936.This model is based on the kinetic equation of first-order and it is used to this day in modifications for the watching of the cellulose degradation in different conditions. Hill et al. (1995 ) deduced a similar model with the

Alterations of DP (degree of polymerisation) of cellulose fibres due to recycling and ageing at the pulp fibres drying temperature of 80°C, 100°C a 120°C.

contribution of the zero order kinetic. Experimental results are often controversial and new kinetic model for explanation of cellulose degradation at various conditions was proposed ( Calvini et al. 2008 ). The first-order kinetic model developed by these authors suggests that the kinetics of cellulose degradation depends upon the mode of ageing. An autoretardant path is followed during either acid hydrolysis in aqueous suspensions or oven ageing, while the production of volatile acid compounds trapped during the degradation in sealed environments primes an autocatalytic mechanism. Both these mechanisms are depleted by the consumption of the glycosidic bonds in the amorphous regions of cellulose until the levelling-off DP (LODP) is reached.

At the accelerated ageing ofnewspaper ( Kačík et al. 2008 ), the cellulose degradation causes the decrease of the average degree of polymerisation(DP). The DP decrease is caused by two factors in accordance with equation

DP = LODP + DP01.e -k1.t + DP02.e -k2.t ,

where LODP is levelling-off degree of polymerisation. There is a first factor higher and quickdecreasing during eight days and a second factor is lower and slow decreasing and dominant aftereight days of the accelerating ageing in the equation. The number of cleavaged bonds can be welldescribed by equation

DP 0 /DP t – 1 = n 0 .(1-e -k.t ),

where n 0 is an initial number of bonds available for degradation. The equation of the regression function is in accordance with Calvini et al. (2007 ) proposal, the calculated value (4.4976) is in a good accordance with the experimentally obtained average values of DP 0 a DP 60 (4.5057). The DP decreased to cca 38% of the initial value and the polydispersity degree to 66% of the initial value. The decrease of the rate constant with the time of ageing was obtained also by next authors ( Emsley et al. 1997 ; Zervos & Moropoulou 2005 ; Ding & Wang 2007 ). Čabalová et al. (2011 ) observed the influence of the accelerated ageing on the recycled pulp fibres, they determined the lowest decrease of DP at the fibres dried at the temperature of 120°C ( Fig. 10 ).

The simultaneous influence of the recycling and ageing has the similar impact at the drying temperatures of 80°C (decrease about 27,5 %) and 100°C (decrease about 27.6%) in regard of virgin pulp, lower alterations were at the temperature of 120°C (decrease about 21.5%). The ageing of the recycled paper causes the decrease of the pulp fiber DP, but the paper remains good properties.

4. Conclusion

The recycling is a necessity of this civilisation. The paper manufacturing is from its beginning affiliated with the recycling, because the paper was primarily manufactured from the 100 % furnish of rag. It is increasingly assented the trend of the recycled fibers use from the European and world criterion. The present European papermaking industry is based on the recycling.

The presence of the secondary fibres from the waste paper, their quality and amount is various in the time intervals, the seasons and the regional conditions. It depends on the manufacturing conditions in the paper making industry of the country.

At present the recycling is understood in larger sense than the material recycling, which has a big importance from view point of the paper recycling. Repeatedly used fibres do not fully regenerate their properties, so they cannot be recycled ad anfinitum. It allows to use the alternative possibilities of the paper utilisation in the building industry, at the soil reclamation, it the agriculture, in the power industry.

The most important aim is, however, the recycled paper utilisation for the paper manufacturing.

Acknowledgments

This work was financed by the Slovak Grant Agency VEGA (project number 1/0490/09).

• 11. CEPI (Confederation of European Paper Industries). 2006 Special Recycling 2005 Statistics- European Paper Industry Hits New Record in Recycling. 27.02.2011, Available from: http://www.erpa.info/images/Special_Recycling_2005_statistics.pdf
• 12. CEPI (Confederation of European Paper Industrie). 2010 Annual Statistic 2009. 27.02.2011, Available from: http://www.erpa.info/download/CEPI_annual_statistics%202009.pdf
• 18. European Declaration on Paper Recycling 2006 2010 , Monitoring Report 2009 (2010), 27.02. 2011, Available from: http://www.erpa.info/images/monitoring_report_2009.pdf

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A Systematic Review of E-Waste Generation and Environmental Management of Asia Pacific Countries

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Due to the rapid increase in the use of electrical and electronic equipment (EEE) worldwide, e-waste has become a critical environmental issue for many governments around the world. Several studies have pointed out that failure to adopt appropriate recycling practices for e-waste may cause environmental disasters and health concerns to humans due to the presence of hazardous materials. This warrants the need for a review of the existing processes of e-waste management. In view of the growing e-waste generation in the Asia Pacific region and the importance of e-waste management, this study critically reviews previous research on e-waste generation and management practices of major e-waste producing nations (Australia, China, India, Indonesia, and Malaysia) in the Asia Pacific region, provides an overview of progress made and identifies areas for improvement. To fulfil the aims of this research, previous studies from 2005 to 2020 are collected from various databases. Accordingly, this study focuses on e-waste generation and environmental management of these countries. This study found that e-waste management practices of the selected countries need to be enhanced and recommends several best practices for effectively managing e-waste.

1. Introduction

The Asia Pacific region is highly populated and is considered one of the fastest developing regions in the world. In addition, many countries in this region underwent rapid industrialisation, driven by foreign direct investments [ 1 ] due to a relatively cheap labour force. One of the industries that benefited from these factors is the electrical and electronics industry, which has experienced a major transformation due to increased technological and market developments [ 2 ]. Today, electrical and electronic equipment (EEE) has become indispensable and enhance living standards, but often contain toxic chemicals that negatively impact human health and the environment and fuel the climate crisis [ 2 , 3 ]. The growth in demand and increased sales of EEE have consequently led to the rise in the volume of e-waste [ 3 , 4 , 5 ].

E-waste is one of the most urgent and pressing challenges of our time; however, it is routinely ignored. Across the world, the growing amount of e-waste threatens the environment and local communities, as incorrectly disposed e-waste results in life-endangering toxic chemicals released into the environment and the loss of precious metals [ 2 , 4 , 5 , 6 , 7 ]. Perkins et al. [ 8 ] point out that the amount of e-waste generated each year is increasing at an alarming rate. In 2019 alone, more than 50 million tons (Mt) of e-waste was generated globally. Of this total e-waste, 24.9 million tons were generated in the Asia Pacific region alone. The amount of e-waste generated worldwide increased three times faster than the world’s population. Forti et al. [ 2 ] estimate that the volume of e-waste generated globally will exceed 74 million tons (Mt) by 2030. However, the level of recycling is not keeping up the pace. In fact, less than 13 per cent of e-waste was recycled in the same year. Moreover, the majority of e-waste generated is being diverted for landfilling, which is a common approach to disposing of e-waste worldwide [ 9 ]. The major issue with the current e-waste management practices is: (a) lack of efficient collection and recycling systems and (b) lack of mechanisms to hold producers of EEE accountable for the end-of-life disposal [ 2 ]. Hence, failure to adopt appropriate e-waste recycling processes may lead to enormous environmental and health issues [ 3 , 10 , 11 , 12 , 13 ].

This study identified three research gaps. Firstly, although, literature presents results of various studies on e-waste generation [ 3 , 4 , 5 , 8 , 14 , 15 , 16 , 17 ], recycling [ 14 , 15 , 16 , 17 ], treatment [ 4 , 18 , 19 , 20 ], and environmental management [ 8 , 21 , 22 , 23 , 24 ]; however, few studies have focused on the impact of e-waste generated in the Asia Pacific countries selected and its consequential effects on human health and the environment. Secondly, Forti et al. [ 2 ] suggest that many countries, including countries in the Asia Pacific region, are not sufficiently managing e-waste generated, and greater effort is needed to ensure smarter and more sustainable global production, consumption, management, and disposal of e-waste. The authors also indicated that more e-waste is generated than is being safely recycled in many countries of the world, and more corporative efforts are needed to tackle the escalating e-waste problem through appropriate research and training. Forti et al. [ 2 ] and Balde et al. [ 3 ] noted that the issues emanating from e-waste management in today’s digitally connected world are primarily due to the way we produce, use, and dispose of electronic devices, which are currently unsustainable. Bhaskar and Kumar [ 25 ] added that implementing appropriate e-waste management strategies will contribute to the achievement of sustainable development goals and reduce the global climate crisis through developing the necessary, needed, and required e-waste policies. Thirdly, while investigations and discussions on e-waste generation and management have been ongoing for several decades. However, the problems and challenges on e-waste generation and management remain unabated [ 2 , 26 , 27 ].

The purpose of this study is to critically review the existing strategies and practices adopted by the major e-waste producing countries in the Asia Pacific region in managing and regulating e-waste to minimise the environmental and health impacts created as a result of inappropriate recycling and disposal practices.

A key initiative and motivation of this study is to identify the problems/challenges in managing e-waste in the selected Asia Pacific countries and recommend appropriate management strategies and policy approaches to handle and regulate e-waste to significantly reduce environmental and health concerns. Accordingly, this study reviews previous research on e-waste generation and environmental management of Australia, China, India, Indonesia, and Malaysia, identifies problems and challenges that negatively impact e-waste management in these countries, provides an overview of progress made, and identifies areas for future research.

The selected countries (Australia, China, India, Indonesia, and Malaysia) are among the largest producers of e-waste in the Asia Pacific region [ 2 , 13 , 18 , 28 ]. To fulfil the aims of this study, a comprehensive review of previous research articles on e-waste published from 2005 to 2020 was conducted. This study focuses on aspects such as the amount of e-waste generated, current recycling and disposal methods, environmental management of e-waste, individual/collective attitudes towards e-waste, current e-waste problems/challenges of selected countries. In addition, prior studies of the selected countries are categorised based on the type and scope of research, location of study, and e-waste categories analysed. This study uses the outcomes of previous studies, considers country-specific issues, and identifies future research areas to present best practices for e-waste generation and environmental management.

This paper is organised into five sections. The first section presents current literature on e-waste, the research problem, research gaps and research aim, and justification for this study. The second section outlines the chosen methodology and the justification for considering a systematic literature review. The third section details the e-waste management practices in the selected countries. The fourth section provides the results of this study and analyzes the results. The final section presents the findings of this study, limitations associated with the current study, policy recommendations for effective e-waste management, and future research opportunities.

2. Research Methods

In recent years, researchers have increasingly used quantitative and qualitative research (mixed methods) techniques to expand the scope and improve the analytic power of their studies [ 29 , 30 ]. Quantitative research method is a statistical and interpretive technique used to describe or explain the meaning and relationships of a phenomenon under investigation. Quantitative research typically involves probability sampling to allow statistical inferences to be made [ 29 , 31 ]. In contrast, qualitative research method is a non-numerical, precise count of some behaviour, attitudes, knowledge, or opinion for ascertaining and understanding the meaning and relationships of certain phenomena for generalisation. It typically involves purposeful sampling to improve understanding of the issues being examined [ 29 , 30 , 31 ].

This study adopts a qualitative research method to explore the issues relating to e-waste in the selected countries from existing research over the past years to guide future research in this area. To achieve the aim of this study, the five-phase approach of Wolfswinkel et al. [ 32 ] for conducting a systematic review and analysis of the literature is adopted. Adopting this five-phase approach enables the researchers to conduct a thorough search process and critically review and analyse the articles retrieved from the databases. The five-phase approach includes: (a) defining the scope of the review, (b) searching the literature, (c) selecting the final samples, (d) analysing the samples using content analysis, and (e) presenting the findings.

The first phase is to define the scope of the review. This includes the definition of specific criteria for the inclusion and exclusion of relevant sources and the criteria for identifying and retrieving those sources in the literature. In this study, four prominent databases are used to source literature, including ProQuest, Emerald, ScienceDirect, and Web of Science. The selection of these databases is due to their representativeness and coverage in the publication of top academic papers on e-waste in the selected countries. To ensure broad coverage of the studies in these databases, several keywords have been used for the search, which includes “electronic waste”, “e-waste”, “waste electrical and electronic equipment”, “e-waste management”, “e-waste recycling,” “e-waste disposal methods”, “e-waste problems and challenges” and “environmental management of e-waste”. Several criteria are used to set the limitation, including restricting the document type to scholarly journals, peer-reviewed conference papers, book chapters, and other institutional reports from United Nations (UN) and World Health Organization (WHO); the language in English, and the publication date from 2005 to 2020. These document types have been selected as they represent state-of-the-art research outputs with high impact [ 32 ].

The second phase is to run the search query within the selected databases for retrieving the search results. A total of 688 articles are returned using the above pre-defined search strings. This initial search enables us to gain a general understanding of the coverage of e-waste topics.

The third phase involves selecting the final samples for detailed analysis. The search is limited to the title and the abstract to focus on the search results. Titles and abstracts of all initial articles are screened for checking the relevance to e-waste. This leads to the identification of 235 relevant articles. Duplicate articles are removed. A total of 210 articles is assessed for eligibility, and after excluding those articles that did not meet eligibility criteria, a total of 185 articles is identified for further review.

The 185 articles have been read in full for coding and analysis. NVivo 12.0 is used for providing an overview of the general topics from all the abstracts of the included papers. An overview of the dispersion of the selected papers in terms of year of publication shows there is increased interest in e-waste from 2005 to 2020. Figure 1 below illustrates the search process using the PRISMA flow diagram.

PRISMA flow chart indicating the results of searches.

3. Overview of E-Waste

E-waste is defined as an electrical appliance that no longer satisfies the user for its intended purpose [ 33 ]. Meanwhile, StEP [ 34 ] defines e-waste as a term used to cover items of all types of EEE and its parts that have been discarded by the owner as waste without the intention of reuse.

Table 1 shows e-waste generated around the world and per continent in 2016. It is observed that the Asian continent generated the highest e-waste, followed by Europe and the Americas. Interestingly, the African continent produced one of the lowest e-waste even though it is the second most populated continent in the world [ 35 ]. Although the African continent produced the lowest amounts of e-waste due to slow technological growth and limited access to energy when compared to other continents, they suffer other kinds of pollution problems caused by traffic emissions, oil spills, heavy metals, refuse dumps, dust, and open burnings and incineration, which significantly contribute to environmental contamination in Africa [ 36 , 37 , 38 ]. Human exposure to toxic metals and environmental pollution has become a major health risk in Africa and is the subject of increasing attention to national and international researchers and environmentalists [ 37 , 38 ].

E-waste generated around the world and per continent in 2016 [ 4 ].

A further study was conducted in 2019 whereby the Asia Pacific region also generated the highest amount of e-waste in comparison to America, Europe, Africa, and Oceania regions. The Asia Pacific region generated around 25 Mt, followed by America at 13.1 Mt and Europe at 12.1 Mt. The study also showed that Africa generated 2.9 Mt and Oceania generated 0.7 Mt of e-waste [ 2 , 39 ]. This warrants the need to conduct a study on e-waste generation and environmental management of countries in the Asia Pacific region [ 14 , 15 , 40 ].

3.1. Constituents of E-Waste

Over the years, the use of electronic devices for domestic and commercial purposes has grown rapidly [ 8 ]. E-waste generally consists of a range of hazardous materials ( Table 2 ), including metals, pollutants, printed circuit boards, computer monitors, cables, plastics, and metal-plastic mixtures [ 2 ]. The composition and quantities of these materials vary in each electronic device depending on the manufacturer, the equipment type, model, and the age it was discarded. In comparison to household e-waste, the e-waste from the IT and telecommunication sector generally contains metals that are of high economic value [ 41 , 42 ]. These metals are generally categorised into precious and toxic metals. Precious metals include gold, silver, aluminium, iron, copper, platinum, etc. The value of precious metals in e-waste is estimated to be worth USD 14 billion. However, more than 50 per cent of these metals are not recovered [ 2 ]. Meanwhile, toxic metals in e-waste include mercury, cadmium, lead, and chromium [ 2 , 43 ].

The distinctive contents of e-waste.

3.2. E-Waste Generation and Management Practices

This study has selected five countries, including Australia, China, India, Indonesia, and Malaysia, from the Asia Pacific region because they are the major e-waste producers in the region. In line with the aim of this study, this section presents an in-depth analysis of waste generation, policies and management practices adopted by the selected countries in the Asia Pacific region. In addition, this section presents literature on e-waste generation and the opinions of scholars in this field. The following sub-sections explain e-waste management practices for the selected countries in the Asia Pacific region. Table 3 below presents e-waste key statistics for the selected countries.

E-waste key statistics 2019.

3.2.1. Australia

Australia is placed among the top 10 consumers of electronic products in the world. As a result, e-waste has become one of the fastest-growing waste streams in Australia [ 9 , 44 , 45 ]. The total and per capita e-waste generation in Australia has steadily increased in the last 10 years from 410 Kilotons (Kt) in 2010 to 554 Kt in 2019 as a result of an increase in sales of EEE [ 2 ]. Previously, due to the lack of an e-waste national regulatory framework, local government councils had difficulties in managing e-waste, and they had no strategies to address e-waste issues [ 46 , 47 ]. To resolve the nation’s escalating e-waste challenges, the Australian government established the National Waste Policy in 2019 to integrate existing policies and regulatory frameworks for e-waste management [ 9 , 45 , 48 ]. Thereafter, the Australian government introduced the National Product Stewardship Scheme in 2011 in collaboration with the State and Territory Governments and industries [ 9 , 26 , 45 ].

The introduction of the National Waste Policy in 2009 was designed to set the direction of Australia’s e-waste management and resource recovery for 10 years from 2010 to 2020. The policy was established to achieve several goals, including compliance to international obligations such as the Basel and Stockholm Conventions, reducing the generation of e-waste, and ensuring e-waste treatment, disposal, recovery, and reuse is safe and environmentally sound [ 44 , 47 ]. The Product Stewardship Act of 2011 was also designed to establish a framework by which the environmental, health, and safety impacts of electrical and electronic equipment and its recycling and disposal are adequately managed [ 44 , 45 ]. Currently, Australia’s e-waste system is in its evolving stages and while, progress has been made since the introduction of the National Waste Policy and the Product Stewardship Act, Australia’s e-waste is growing three times faster than other waste streams, and the capacity and sophistication of the nation’s systems need to grow and adapt [ 44 , 48 ].

3.2.2. China

China is one of the leading producers of EEE, and currently, the country is experiencing incredible growth in e-waste generation from both domestic and international sources [ 9 , 26 , 49 ]. Formal e-waste management in China is driven by government agencies designed to improve e-waste recycling and disposal and to encourage manufacturers to take back their products [ 21 , 49 ]. Thus, Chinese e-waste regulations are focused on extended producer responsibility (EPR), polluter pays, and 3Rs (reduce, reuse, recycle) principles [ 50 ].

Informal e-waste recycling in China is often carried out by individual recyclers and unauthorised dismantling companies. Informal recyclers purchase used items and often either dismantle or repair them for the second-hand market. This unregulated e-waste recycling method is currently flourishing in China. Informal recycling provides livelihoods for many Chinese citizens and is creating serious environmental and health concerns. Thus, e-waste generation and management in China has remained a major problem and are fuelled by China’s inexpensive labour and manufacturing abilities. Informal recyclers do the majority of e-waste collection and recycling in most cities throughout China [ 50 ].

3.2.3. India

The increasing average annual growth rate from 0.56% in 1991 to 1.62% in 2011 has contributed significantly to an alarming amount of e-waste generation in India. India is among the top 10 countries in the world in e-waste generation after the U.S. and China. It is estimated that three (3) million tons of e-waste were produced in 2018 and is expected to reach five (5) million tons by the end of 2020 [ 51 , 52 , 53 ]. According to the Confederation of Indian Industries, the Indian electronics industry has a market size of approximately USD 65 billion in 2013, and this is expected to reach USD 400 billion by the end of 2020 [ 52 , 54 ].

Today, e-waste in India is a significant waste stream both in terms of volume and toxicity [ 55 ]. Approximately 152 million units of computers will become obsolete in India by the end of 2021 [ 55 , 56 ], creating serious management challenges and environmental/health problems. Each year, India domestically produces approximately 400,000 tons of e-waste [ 24 ]. Thus, India’s e-waste recycling is a market-driven industry [ 55 ] and is dominated by a number of informal actors. About 90% of the e-waste in India is illegally recycled in the informal sector and involves different groups, including women and children [ 57 , 58 ].

The Ministry of Environment and Forests (MoEF) is the national regulator responsible for formulating legislation related to e-waste management and environmental protection. MoEF approves the guidelines for the identification of the various sources of e-waste in India and endorses the procedures for handling e-waste in an appropriate and environmentally friendly manner [ 59 ]. Those involving e-waste are the 2004 “Municipal Solid Waste Management Rules” and the 2008 “Hazardous and Waste Management Rules.” New regulations are classified as the 2010 “E-waste Management and Handling Rules”, which became effective in 2012 [ 60 ]. While there are regulations on e-waste management and disposal in India, no regulation has effectively addressed the e-waste problem in India [ 52 , 58 ]. Currently, the majority of the hazardous materials found in e-waste are covered under “The Hazardous and Waste Management Rules, 2011 and the 2016 E-waste Management and Handling Rules” [ 52 ].

Despite EPR being a major policy approach in both e-waste (Management and Handling) Rules 2011 and E-waste (Management and Handling) Rules 2016, they are not effectively implemented, and this can be attributed to certain peculiarities in India’s e-waste management system [ 51 , 61 ]. For example, due to some financial incentives involved, Indian consumers are willing to sell their obsolete e-waste to the “kawariwalas” (door-to-door scrap collectors). This behaviour is totally different from practices adopted by most developed countries whereby the producers and consumers have to pay “Recycling/Disposal Fee” [ 62 , 63 , 64 ].

3.2.4. Indonesia

Due to substantial growth in the economy coupled with rapid technological developments, e-waste generation in Indonesia has increased considerably [ 28 , 65 ]. In 2016, Indonesia generated 1274 kt of e-waste with a per capita generation of 4.9 kg [ 66 ]. Although e-waste appears as a global issue, it is not a common term for most people in Indonesia [ 67 , 68 ]. In Indonesia, e-waste management is dominated by the informal recycling sector, which is essentially made of unregulated and unregistered small businesses, groups, and individuals, while the formal sector consists of the country’s municipal agencies as the major actors [ 69 ].

Although the country has no presence of a specific regulation to manage its e-waste, the “Environmental Protection and Management Act No. 32/2009” and “Solid Waste Management Act No. 18/1999” are used in the regulation of e-waste produced in the country [ 70 , 71 ]. Since 2016, the Indonesian government has been in the process of formulating a unified e-waste regulation for the country, which would apply to all the 37 Indonesian provinces, but this is yet to be realised [ 28 , 72 ]. However, the absence of regulated licensed recycling companies in the country has encouraged inappropriate disposal of the majority of the EEE from households, businesses, and industries [ 71 ]. Currently, the informal sector illegally collects, treats, and disposes of discarded EEE triggering huge environmental and health concerns [ 65 , 72 ]. The difficulties/challenges in managing e-waste in Indonesia is primarily due to (a) the inability of the government to understand and deal with the interest of stakeholders involved, (b) the government regulations are beneficial to only a few parties, and (c) there is strong resistance between the government agencies [ 73 ].

3.2.5. Malaysia

In 2019, the International Monetary Fund (IMF), in its economic outlook, ranked Malaysia as the 3rd largest economy in Southeast Asia and the 37th largest economy in the world [ 74 ]. With a healthy economic indicator, e-waste generation in Malaysia is expected to increase in the coming years. The growth in e-waste generation is anticipated worldwide because there is a strong correlation between economic growth and e-waste generation [ 75 , 76 ].

Management of e-waste in Malaysia is still in its infancy and only began in 2005 [ 77 ]. In Malaysia, e-waste is classified as scheduled waste under the code SW 110, “Environmental Quality Regulations 2005” and managed by the Department of Environment (DOE) and the Ministry of Natural Resources and Environment (MNRE) [ 78 , 79 ]. The primary role of DOE and MNRE is pollution prevention and control through the enforcement of the “Environmental Quality Act 1974” (EQA 1974) [ 79 , 80 ]. Although there are strategies on e-waste management in place, they do not adequately guide the local consumers or the municipal authorities on how e-waste should be managed, reused, recycled, or disposed of [ 78 ]. Subsequent to the listing as e-waste under the “Environmental Quality Scheduled Waste Regulations (EQSWR) 2005”, e-waste in Malaysia was reported and managed as municipal solid waste through the Department of Solid Waste Management (DSWM) under the Ministry of Housing and Local Government [ 78 , 81 , 82 ].

3.3. A Review of Previous Studies

This study considered literature reviews to identify key issues associated with e-waste management and to conduct an extensive evaluation of e-waste management practices in the selected countries. We believe this knowledge will help the countries to overcome their challenges and develop appropriate strategies for recycling and disposing of e-waste. This section provides an overview of earlier studies in the selected countries. In particular, results from the literature review on e-waste generation and management practices adopted by the respective nations are presented. Furthermore, this section presents the scope and the context of earlier studies on e-waste management. Prior studies [ 83 , 84 , 85 , 86 ] offer valuable insights into e-waste management in the selected countries. They also highlight the challenges associated with e-waste management and the need for developing comprehensive e-waste management strategies. Table 4 presents previous research on e-waste conducted in the selected countries from 2005 to 2020.

Previous studies on e-waste conducted in the selected countries from 2005 to 2020.

4. Results and Discussion

This study adopts a qualitative approach for studying e-waste management practices of the selected countries in the Asia Pacific region. As per Wolfswinkel et al. [ 32 ], this study adopted a five-phase approach. In the first phase, secondary data from 2005 to 2020 has been considered for reviewing existing literature on e-waste management in the selected countries. Then, a total of eight (8) keywords are used to identify and analyse the relevant articles. Finally, challenges and practices associated with e-waste management are discussed to present the proposed policy approaches and recommendations.

E-waste management has become a contentious issue due to the presence of hazardous materials and the health hazards it may cause if not managed properly. In fact, for more than a decade, scholars have conducted studies on informal e-waste collection and disposal methods [ 87 , 88 ]. However, these studies were limited to e-waste generation, prevention, quantification, recycling, treatment, reuse, pollution control, legislation, and life-cycle assessment, as noted in recent studies [ 83 , 85 , 87 , 89 , 90 , 91 ]. Undoubtedly, these studies presented opportunities to address some of the challenges associated with e-waste management. However, there is a limited study in addressing the environmental and health implications associated with e-waste for achieving sustainable e-waste management. Moreover, prior studies on e-waste are centred on a small number of developed countries, which represent a “standard” or “benchmark” for developing e-waste management policies for emerging countries. Therefore, this study aims to address these gaps.

4.1. E-Waste Studies in Selected Countries

After a critical review of the pertinent literature and a content analysis of the e-waste articles related to the selected countries, the dispersion of e-waste research in the selected countries according to the keywords/themes, e-waste categories examined, and the study location are illustrated in Table 5 . Based on the information presented in Table 5 , it is evident that most of the e-waste studies in the selected countries were focused on e-waste generation, management and recycling. A number of e-waste studies focused on problems and challenges, environmental management, and health impacts indicating that further research is required in these areas in the countries examined.

Distribution of e-waste research in selected countries.

4.2. Analysis of Content Results

Given the background review and analysis in the previous sections, it is obvious that the problem and challenges of e-waste in the selected countries still persist. Our analysis shows that the e-waste management systems and infrastructure of the selected countries, particularly India, China, Malaysia, and Indonesia, are still in their infancy. Currently, e-waste scrap such as printed circuit boards, CRT monitors, and LCD screens have been, and are still being, recycled in China, India, Indonesia, and Malaysia, creating huge environmental and health issues. Informal e-waste collection, recycling, and its health implications on informal workers in these countries have become increasingly popular in the last 15 years [ 89 , 92 , 93 , 94 ]. Table 6 shows the findings from the analysis of the contents.

Findings from the analysis of the contents.

In China, several towns have remained as a dumping ground for e-waste. For example, Guiyu town is often referred to as “the e-waste capital of the world” and employs more than 150,000 locals from four villages. These local informal workers dismantle and recapture valuable metals and parts that can be reused or sold from old computers. In Guiyu, it is not uncommon to see computer parts, cables, and huge tangles of wires scattered around the streets and riverbanks [ 88 , 95 , 96 , 97 ]. Findings/outcomes indicate that various issues geared towards developing a sustainable recycling system still need to be addressed.

In India, obsolete computers from households and businesses are sold by auction to door-to-door collectors who engage in informal methods of recycling. According to a report by the Confederation of Indian Industries (CII), approximately 146,000 tons of obsolete EEE are generated in India annually [ 86 , 109 ]. The results of the analysis show that the recycling of e-waste in India is heavily dominated by the informal sector, and only a few approved e-waste recycling facilities are available. In the majority of urban slums of India, more than 95% of e-waste is treated and processed by untrained workers who carry out illegal and risky procedures. These illegal procedures are not only injurious to the health of the locals who work without personal protective equipment but also to the environment [ 55 , 86 ]. It is found that the formal process of e-waste recycling and treatment is still rather slow as the collection and recycling of most e-waste remains in the hands of the informal sector [ 86 , 109 ].

In Indonesia, large amounts of e-waste are imported from developed countries. E-waste in the form of scrap materials or second-hand devices is sent to Indonesian islands from the adjacent ports in Singapore and Malaysia. Findings indicate that, in Indonesia, infrastructure and workable systems to quantify, recycle, monitor, and handle e-waste is lacking [ 65 , 127 ]. Currently, the informal sector illegally collects, treats, and disposes of discarded EEE, causing huge environmental and health issues [ 65 , 71 ].

The management of e-waste in Malaysia is still developing and only began in 2005 [ 77 ]. Results indicate that although there are strategies to manage e-waste in Malaysia, challenges persist and the pressure to manage e-waste is now even more crucial. Malaysia has become one of the popular destinations of e-waste imported from developed countries [ 139 , 140 , 141 ]. Results of the analysis also indicate the country still faces significant issues in managing the ever-increasing amount of e-waste generated even though several material recovery facilities (MFR) have been established.

In Australia, several government policies have been developed. The key issues are identified in the e-waste management including: (a) the narrow scope of e-waste categories for recycling, (b) the lack of clarity on the roles of key stakeholders involved, (c) the recycling and material recovery targets, and (d) the lack of auditing and compliance. The results of the analysis show [ 47 , 142 , 143 ] minimal research has been undertaken to assess the effectiveness of e-waste policy management strategies [ 47 , 144 , 145 , 146 , 147 ].

It can be seen that the majority of the selected countries in this present study are faced with an increasing amount of e-waste. Although the per capita e-waste generated in the emerging countries is much lesser than in the developing countries, the volume generated is greater due to the growing population and market size in emerging countries such as India, China, and Indonesia. These countries are ranked among the top e-waste generators in the world.

The importance of selecting these countries such as Australia, India, China, Indonesia, and Malaysia in the Asia Pacific region in terms of environmental and market perspectives cannot be overemphasised. These selected countries have significant population, natural resources, and financial potentials [ 67 , 148 , 149 , 150 , 151 ]. Moreover, these countries have contributed substantially to the world’s GDP, landmass, and market share. This calls for a responsible e-waste management effort by these countries to effectively manage the growing amounts of e-waste generated for reducing environmental and health concerns.

Clearly, e-waste management processes in the majority of these countries examined still need improvement. Most of these countries studied have no well-established e-waste infrastructure for efficient collection, storage, transportation, recycling, and disposal of e-waste. In addition, the enforcement of codes of practice and regulations relating to hazardous e-waste management in these countries is minimal or non-existent.

Exposure to e-waste is harmful to public health. E-waste has been found to negatively impact public health because communities are exposed to a complex mixture of chemicals from multiple sources and through multiple exposure routes [ 152 ]. The results of this study indicate that the impact of e-waste is linked to a variety of health problems in the countries examined, such as birth defects, premature births, respiratory diseases, and cancer. Furthermore, people living in e-waste recycling towns or working in e-waste recycling sites showed evidence of greater DNA damage. A review of the literature also revealed an association between e-waste exposure and thyroid dysfunction, adverse behavioural changes, and damage to the lungs, heart, and spleen due to prolonged exposure [ 152 , 153 ].

Hence, e-waste has become one of the major challenges in these countries, and it is, therefore, crucial for these countries to investigate the development of a well-organised and inexpensive recycling scheme to extract valuable resources with inconsequential environmental impacts.

5. Conclusions

This study has evaluated the e-waste generation and management practices of the selected countries in the Asia Pacific region. Based on the review of past studies and results of the analysis, it is obvious that the majority of the selected countries are yet to find a workable e-waste management strategy that will provide a sustainable solution to their e-waste concerns.

Results of the analysis show that the volumes of e-waste generated are fast exceeding the available infrastructure and recycling facilities in the countries examined, thereby driving e-waste streams to flow into illegal and informal recovery. On top of that, the absence of an integrated framework that could support the monitoring and management of toxic and hazardous wastes has also created additional problems in managing e-waste in the selected countries and calls for a generic e-waste policy approach.

In addition, the increasing demand for second-hand EEE, particularly in developing countries (China, Indonesia, India, and Malaysia) due to poverty and the continuing technological modernisation, has made these countries dumping grounds for e-waste from developed countries. For example, China’s Guiyu town is well-known for the informal recycling of printed circuit boards. Specifically, “metal-contaminated sediments and elevated levels of dissolved metals have been reported in rivers around the town of Guiyu” [ 85 ].

Furthermore, sophisticated facilities and infrastructure required for formal recycling of e-waste using efficient technologies are minimal or non-existent in the selected countries. Formal recycling is widely accepted as the best way to manage e-waste, which reduces greenhouse gas emissions and helps lessen the climate crisis. Thus, recycling e-waste will reduce air and water pollution associated with the illegal dumping of e-waste. By recycling discarded, unwanted, or obsolete EEE for new products, nations can further reduce the enormous health risks and environmental pollution associated with improper disposal of e-waste.

Therefore, to effectively manage e-waste in the selected countries, there is a need to develop generic structured policy approaches to tackle the e-waste problem in the selected countries and indeed across the world is required. These structured policies are projected to put in place formal systems and infrastructure for the recycling, management, and disposal of e-waste, taking into account country-specific issues.

One of the shortcomings of this study is that the information and analysis of previous studies are seen to be reality. This study is also limited to countries in the Asia Pacific region and considers the time limitation by the year of the articles found. Although the accuracy of some of the analyses in the present study is inescapably subjective, this study is a starting point for further research into various aspects of e-waste generation and management practices of the selected countries.

6. Recommendations

This study has exposed the current situation of e-waste generation and management practices of the selected countries. The following recommendations are suggested based on the findings of this study:

• E-waste regulations tailored to each country’s current situations should be enacted, recognising the lessons learned from more developed and experienced nations such as Japan, Switzerland, and South Korea;
• Extended producer responsibility (EPR) and 3Rs strategy should be implemented in EEE manufacturing regulations in all countries to support the production of simple, lightweight products, planned for reuse rather than obsolescence so that recycled materials can become resources for new products, thereby reducing the request for raw materials;
• Local government councils are key stakeholders in the management and recycling process and therefore incur major expenditures while handling e-waste. This, therefore, necessitates policymakers understanding of the determinants, drivers, and costs associated with e-waste collection and disposal;
• International integrated organisations should be established for checking specific e-waste material generation across the globe. This initiative will restrain the transboundary movement of e-waste across international borders.

Policy Approaches

Although different countries have endorsed and passed their respective e-waste regulations in other to manage e-waste, implementing appropriate and structured policy approaches will support all efforts directed towards effectively managing e-waste across the globe. Firstly, it is critical to have stepwise, and well-thought-out policy approaches for effectively formulating and implementing e-waste regulations and guidelines. Such approaches have been found to be effective in more advanced countries such as Switzerland, South Korea, and Japan, as noted above. In view of the multidimensional socio-economic nature of emerging economies, it is vital to consistently assess and evaluate existing policies to identify gaps and areas for improvement. This technique has also been found to be effective in Australia. Secondly, when implementing e-waste policies, interdisciplinary research approaches need to be considered. This will allow policymakers to better understand and address the various health and environmental problems associated with e-waste management. Finally, we believe that the policy approaches of respective countries geared towards dealing with the persistent and challenging e-waste issues require a local and specific approach where inherent socio-cultural, economic, political, and environmental concerns of that country are taken into consideration.

7. Future Research

Future research should use a quantitative approach or other research methods and expand the number of selected countries to understand e-waste generation and management practices of countries in the Asia Pacific region. This will provide additional viewpoints in the management, recycling, and environmental management of e-waste in the regions.

Author Contributions

L.A.: Conceptualisation, Methodology, Formal analysis, Investigation, Resources, Writing—Original Draft; S.W.: Visualisation, Validation, Writing—Review and Editing, Supervision; S.G.: Visualisation, Validation, Writing—Review and Editing, Supervision. All authors have read and agreed to the published version of the manuscript.

This research received no external funding.

Institutional Review Board Statement

Informed consent statement, data availability statement, conflicts of 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.

Publisher’s Note: MDPI stays neutral with regard to jurisdictional claims in published maps and institutional affiliations.

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• Published: 31 August 2024

Economic viability requires higher recycling rates for imported plastic waste than expected

• Kai Li   ORCID: orcid.org/0000-0002-1324-1386 1 ,
• Hauke Ward 1 ,
• Hai Xiang Lin 1 , 2 &
• Arnold Tukker   ORCID: orcid.org/0000-0002-8229-2929 1 , 3  

Nature Communications volume  15 , Article number:  7578 ( 2024 ) Cite this article

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• Environmental economics
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The environmental impact of traded plastic waste hinges on how it is treated. Existing studies often use domestic or scenario-based recycling rates for imported plastic waste, which is problematic due to differences in recyclability and the fact that importers pay for it. We estimate the minimum required recycling rate ( RRR ) needed to break even financially by analysing import prices, recycling costs, and the value of recycled plastics across 22 leading importing countries and four plastic waste types during 2013–2022. Here we show that at least 63% of imported plastic waste must be recycled, surpassing the average domestic recycling rate of 23% by 40 percentage points. This discrepancy suggests that recycled plastics volumes from the global North-to-South trade may be underestimated. The country-specific RRR provided could enhance research and policy efforts to better quantify and mitigate the environmental impact of plastic waste trade.

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Introduction.

Over the past decades, increasing globalisation has fragmented supply chains, making the assessment of life-cycle environmental impacts more challenging 1 , 2 , 3 . A similar trend has emerged in waste management. Since 2019, traded waste plastics have amounted to approximately five million tons per year 4 . Typically, this waste is exported from high-income countries to low-income countries, where labour and treatment costs are lower 5 . However, such exports to the Global South have raised major concerns due to potential mismanagement, which can have severe negative impacts on the environment, ecosystems, and human health 5 , 6 , 7 , 8 . Mismanaged plastic waste contributes to river pollution and is a significant factor in the ‘plastic soup’ found in oceans 9 .

Recent publications indicate that globally less than 10% of waste plastics are recycled 5 . A significant amount of plastic waste is mismanaged in countries with underdeveloped waste collection and treatment systems 10 . For example, over half of the plastic waste in Indonesia is incinerated without recovering energy, and 5% is disposed of in uncontrolled dumpsites 11 . Evidence shows that more than 60% of marine litter plastics emissions annually come from the Philippines, India, Malaysia, and Indonesia 7 . Much of this plastic waste treated in the Global South originates from the Global North, contributing to environmental plastic waste emissions 12 .

In response to these concerns, China, a major importer of plastic waste, implemented a plastic import ban in 2018 8 . This decision redirected plastic waste exports to other countries, notably Malaysia, Indonesia, Turkey and Vietnam. To address potential negative impacts and prevent mismanagement abroad, the European Union (EU) has recently considered a ban on plastic waste exports to non-OECD countries 13 , adhering to the principle that countries should be responsible for the proper treatment of their own waste 14 .

While the global plastic recycling rate remains low, there is an implicit assumption that traded plastic is primarily recycled 4 . However, accurately determining the recycling rate for imported plastic waste in receiving countries is challenging due to measurement difficulties. Existing studies often rely on assumed domestic or scenario-based recycling rates, which lack robust data support. For example, Wen et al. quantified the changes in environmental impacts resulting from the shift of plastic waste imports from China to Southeast Asia, using assumed domestic recycling rates of 10 to 40% for five Southeast Asian countries 8 . Similarly, Bourtsalas et al. estimated the environmental impacts of treating imported plastic waste in the USA, using widely varying recycling rates from 8.7 to 50% 15 . Bishop et al. faced a lack of official data on exported plastics from Europe, leading them to use a broad range of recycling rates from 50 to 90% 16 . This reliance on domestic or scenario-based rates highlights the urgent need for comprehensive and transparent data to guide policy and research effectively.

Moreover, replacing the recycling rate of imported plastic waste with the domestic average is questionable for two main reasons. Firstly, domestic plastic waste often comes from diverse sources, resulting in heterogeneous and difficult-to-recycle mixtures, particularly in regions with inadequate or partial waste separation. In contrast, imported plastic waste is typically more concentrated and uniform, as it is pre-selected for exporting. Secondly, the UN Comtrade database shows that importing countries pay for plastic waste, indicating its economic value (see Fig.  1 ) 17 . If these imports were not processed into valuable recyclates—i.e., if they were primarily dumped or burned—the importing companies would face significant financial losses, making it unsustainable for them. Therefore, any viable approach must ensure that at least part of the imported plastic is converted into economically valuable outputs through recycling to offset initial costs.

The trade value reflects the cost, insurance and freight (CIF) price. The average unit price of waste import is calculated based on the weighted values of four plastic waste types (PE, PS, PVC and others) across years. The original trade data, including net weight and trade value, are sourced from the UN Comtrade database. The volume of plastic waste import is reported in net weight (million tonnes, Mt).

In this work, we introduce a novel approach by defining the Required Recycling Rate ( RRR ). We estimate the RRR for the 22 largest plastic waste-importing countries from 2013 to 2022 based on the economic break-even point, where the revenue from recycling matches the costs of imports and the recycling process (labour, electricity, and real estate rentals 18 , 19 ). We assume that recyclates can be sold at prices comparable to primary plastics and consider physical losses throughout the recycling process 20 , 21 . Import costs and primary plastic values are derived from 186,861 bilateral trade records for four plastic wastes (PE, PS, PVC and others) and six primary plastics (HDPE, LDPE, PS, PVC, PET and PP) from the UN Comtrade database. Here, 'recycling' specifically refers to mechanical recycling, the predominant method for recycling imported waste in Global South countries 22 , 23 . Our findings indicate that the RRR for imported plastic waste in the 22 research countries significantly surpasses their reported national recycling rates. Sensitivity and Monte Carlo-based uncertainty analyses further confirm the robustness of these results.

Required recycling rates across four plastic wastes and 22 countries

Our analysis shows that at least 63% of the imported plastic waste must be recycled to offset the costs. However, the RRR s vary across countries and plastic waste types (Fig.  2 ).

The RRR is displayed for waste PE ( a ), waste PS ( b ), waste PVC ( c ), and waste ‘Others’ ( d ). The 22 research countries are geographically divided into five country groups. For each country, the left bar represents the costs associated with recycling 1 kg of plastic (including plastic waste imports), while the right bar shows the value of 1 kg of recycled plastic. The RRR calculated using mirror trade data is shown in Supplementary Fig.  1 , with comparisons detailed in Supplementary Table  1 . The annual RRR from 2013 to 2022 across 22 research countries and four plastic waste types is presented in Supplementary Figs.  2 – 5 .

Due to the significant gap between recycling costs and product prices, countries in Asia and Eastern Europe have the lowest RRR for imported plastic waste, starting at around 40%. Specifically, Thailand, Turkey, and the Czech Republic have the lowest RRR benchmarks for their respective plastic waste types, ranging from 40 to 50%. In contrast, higher RRR s are needed for Western Europe and North America, reflecting limited profitability for recycling imported plastic waste in these regions. The highest RRR s are observed in France, the UK, Belgium, and Canada, with average values between 61% and 82% for all plastic waste types. For comparison, the mechanical recycling costs collected from other literature sources are detailed in Supplementary Table  2 .

Examining distributions among four plastic waste types reveals that PE and PS waste have the lowest RRR , averaging 10–20% lower than those for PVC and ‘Others’ within the same region. PVC recycling is already hindered by challenges such as its chlorine content and contamination issues 24 . Our results further imply that recycling PVC is less economically competitive compared to other plastics, due to a narrower profit margin between recycling costs and primary plastic prices. This results in RRR s for PVC that are on average 14% higher than for PE from 2013 to 2022. The higher RRR s for ‘Others’ plastic waste (Fig.  2d ) are attributed to greater variability in recycling costs and primary plastic prices, with mixed plastic waste falling into this category.

The variation in RRR across different types of plastic waste serves as a crucial market signal for each country’s plastic waste import structure. For example, the Netherlands demonstrates a significant contrast in RRR between waste PVC and waste PE, with RRR of 83% and 62%, respectively. This difference suggests implicitly higher recycling costs and narrower profit margins in the PVC recycling market compared to the PE recycling market in the Netherlands. Confirming this trend, the Netherlands evidenced higher imports of waste PE (3 Mt) compared to waste PVC (0.1 Mt) during the period 2013–2022. Similar import structures are observed in countries such as Germany, the USA, France, and Belgium. In contrast, RRR differences across plastic waste types are less pronounced among countries in the Global South. For instance, RRR s across four plastic waste types range from 50–64% in Vietnam, 40–50% in Turkey, and 50–64% in India. Supplementary Table  1 provides a detailed comparison of RRR differences among plastic waste types across countries.

Although import costs are the largest component of overall expenses and are often seen as a major factor influencing plastic waste trade 25 , they do not fully explain the observed differences in RRR between Europe and Asia as effectively as labour costs do. For instance, Germany, a major plastic waste importer in Western Europe, faces import costs ranging from $0.33 to $0.57 per kilogram across the four plastic types assessed. These costs are only slightly higher than those of large Asian importers such as Turkey ($0.27–$0.53) and Thailand ($0.27–$0.46). In contrast, labour costs, the second largest cost factor, are significantly higher in Western Europe. Recycling 1 kg of imported plastic waste in Germany incurs average labour costs of $0.26 from 2013 to 2022, which is approximately four times higher than in Turkey ($0.067) and five times higher than in Thailand ($0.052). It is important to note that these cost-related statistics, collected at the country level, may not fully capture regional variations within countries.

Comparison between the required recycling rate and the domestic average

A notable discrepancy emerges when comparing the calculated RRR with the collected national plastic recycling rates across 22 countries (see Supplementary Table  3 ). The RRR averages 63% for the period between 2013 and 2022, which is 40% higher than the average national plastic recycling rate of 23% (Fig.  3a ). The RRR average and the domestic average are weighted by import mass across countries and plastic waste types and by the annual domestic plastic waste generated across countries, respectively (refer to Supplementary Table  3 ). This discrepancy subsequently affects the estimation of the amount of plastics recycled from the waste traded from the Global North to the Global South (refer to Fig.  3b , with the countries involved detailed in Supplementary Table  4 ). Using the national plastic recycling rate, the annual amount of recycled plastics from the plastic waste trade from the Global North to the Global South averaged 0.37 million tonnes per year (Mt yr –1 ) over the past five years (2018–2022). In contrast, the annual recycling volume surges to 1.04 Mt yr –1 if the RRR is used, an increase of 0.67 Mt, roughly equivalent to France’s recycled plastics output in 2022 26 .

a Illustrates the variations in average RRR across countries and plastic types. The dashed line represents the average RRR across countries, weighted by the total import mass across countries. The plastic waste type label on the Y axis displays the average RRR of each plastic waste type, weighted by the import mass of each plastic waste type across countries. In addition, the dotted line below denotes the country’s average plastic recycling rate, weighted by the annual domestically generated plastic waste across countries. Mass data corresponding to ( a ) are provided in Supplementary Table  3 . A comparison of the average RRR , weighted by either trade mass or trade value and using both trade and mirror trade data, is shown in Supplementary Table  5 . b Illustrates how discrepancies between these two recycling rates affect the estimates of recycled plastics from the waste traded from the Global North to the Global South between 2013 and 2022 (countries involved are listed in Supplementary Table  4 ). The trade data originate from the UN Comtrade database. The results from ( a , b ) calculated using mirror trade data are shown in Supplementary Fig.  6 .

Assessing uncertainty in the required recycling rate

We conducted a sensitivity analysis to examine how six key variables influence our results by country and plastic waste type (Fig.  4 ). The analysis considers both pessimistic and optimistic scenarios, representing each variable based on their minimum and maximum values observed from 2013 to 2022.

The selected countries include the Netherlands ( a – e ), the USA ( f – j ), Malaysia ( k – o ), the Czech Republic ( p – t ), and Turkey ( u – y ). Sensitivity analysis and Monte Carlo simulation results for other countries are presented in Supplementary Figs.  7 – 15 . The length and colour depth of the horizontal bars represent the range of sensitivity results. The variable ‘product price’ refers to the value of recycled plastics.

On average, fluctuations in the product price for recycled plastics have the greatest impact on the calculated RRR , varying between −25% and +36%. Variations in RRR are also significantly affected by import costs, with changes ranging from −21% to +29%, and physical losses, ranging from −8% to +11%. Labour costs contribute to fluctuations of −4% to +4%, electricity costs from −1% to +2%, and rental costs from −0.9% to +1%. Regional differences are particularly notable in labour costs, with pronounced variations between Europe and Asia. For example, the Netherlands shows fluctuations ranging from −11% to +6% (Fig.  4b–e ), compared to narrower ranges of −4% to +2% in Malaysia and −1% to +1% in Indonesia. In addition, the analysis reveals notable variations across waste types, particularly for PVC and ‘Other’ types. The variations in RRR  of these two waste types are also attributed to fluctuations in product prices and import costs.

The divergent recycling rates across countries result in varied estimations of recycled volumes from the plastic waste trade, complicating the assessment of its environmental impacts. Notably, the RRR averaged ~63% across 22 major importers and four plastic waste types from 2013 to 2022, significantly higher than the average domestic recycling rate of 23%. Moreover, country-specific RRR values exceed those reported in previous studies based on domestic recycling rates. For example, while Wen et al. 8 assumed a recycling rate of 38% for imported plastic waste in Malaysia for 2018, our study indicates a minimum required recycling rate in Malaysia of 58% (PE and PVC), 62% (PS), and 64% (Others) over the period from 2013 to 2022. Such variations in recycling rates can lead to differing estimates of the environmental impacts associated with the plastic waste trade. Higher recycling rates suggest reduced emissions from the avodied virgin plastic production 27 . For instance, Wen et al. 8 assessed the environmental impact of China’s plastic import ban using domestic recycling rates for imported plastics, estimating the net carbon emissions of treating traded plastic waste in 2018 at 0.13 Mt CO 2 -eq. In a scenario reflecting a 50% increase in countries’ recycling rates, closely aligning with our calculated RRR for the same period, this figure dropped to −60 Kt CO 2 -eq.

RRR enhances the accuracy of modelling the fate and impacts of traded plastic waste, which is crucial for scientific research and policy implementation. By indicating the proportion of recycling versus non-recycling, RRR provides valuable data for assessing the environmental impacts of the global plastic waste trade, particularly in waste-importing countries. Moreover, the annual RRR data across countries and plastic waste types sheds light on how external events influence the global plastic waste trade. For instance, a notable increase in RRR across many countries in 2020 coincided with a drop in crude oil prices 28 , suggesting that lower prices for virgin and recycled plastics necessitated a higher RRR to cover costs and achieve profitability. In terms of policy implications, RRR can assist waste-importing countries in formulating and adjusting their recycling targets. Instead of relying on domestic recycling rates, which are often based solely on domestically generated plastic waste, countries should consider separate targets for imported plastic waste, recognising their distinct characteristics.

Our research indicates that while the average RRR of 63% is higher than the domestic average of 23% across 22 research countries, it still falls short of ideal recycling rates. This gap suggests a significant portion of traded plastics may be mismanaged 29 . To address this, transparent tracking systems, such as a robust prior informed consent procedure 30 , are essential. The OECD control system for waste recovery serves as a notable example, requiring disclosure of pre-consented recovery facilities and technologies in waste-importing countries 31 . Although recycling costs may be higher in developed countries, the overall environmental impact is often lower compared to that in Southeast Asia. These environmental concerns are reflected in the EU’s newly adopted waste shipment regulation, which bans plastic waste exports to non-OECD countries starting in November 2026 13 .

Our approach to calculating the RRR focuses on primary cost factors, providing a minimum benchmark for recycling imported plastic waste. The actual RRR might be higher when considering additional costs like environmental costs, capital investment, and operational expenses (e.g., chemical feedstocks 18 , maintenance 32 and value-added taxes 19 ). Although limited cost factors may underestimate the RRR , this method aligns with our goal of establishing a minimum benchmark, providing a better calibration than the domestic recycling rates previously used for imported plastic waste. Future research should explore the full costs and benefits of imported plastic waste for a more comprehensive RRR assessment.

Due to data constraints, we used primary plastic exports as a proxy for recycled plastic revenue to ensure consistency. However, advancements in recycling technologies (e.g., chemical, enzymatic and solvent-based methods) may create higher-value products not captured by current primary plastic classifications 33 , 34 , potentially leading to an overestimation of the RRR in some developed countries. In addition, the four HS codes under 3915 may not fully reflect the quality and diversity of plastic waste, indicating a need for expanded classification coverage.

While our work provides valuable insights into country-specific recycling rates, it does not address regional disparities within countries. Variations in costs such as electricity, labour, and rent can be significant within a country, underscoring the need for future research to determine RRR at regional and city levels for more localised policymaking. Caution is also advised when applying the RRR to estimate recycled volumes and environmental impacts in trade transit countries or regions like Hong Kong (China), as inaccurate trade data may lead to errors in the RRR calculation. Prioritising actual recycling rates through mass balance is recommended for precision, though the RRR remains useful for addressing data gaps and estimating rates where physical measurement is impractical.

Accessing the required recycling rate

Importers aim to profit from recycling plastic waste, but face uncertainty since the recyclability of the waste is often unknown until the container is opened 35 . They must balance maintaining sufficient plastic waste feedstocks to ensure consistent recycling production while keeping both importing and recycling costs below the market value of secondary plastics. The cost factors include expenses related to plasitc waste import, labour, electricity, rent, and physical losses during recycling. We link the recycling rate of imported plastic waste to a cost-benefit inequality (Eqs. ( 1 )–( 3 )) spanning 2013 to 2022, where costs should be less than or equal to benefits. For each year within this period, we selected importing countries that accounted for 70% of the global plastic waste imports, resulting in a total of 22 research countries. This equation can be enhanced with regional data on costs, providing a more accurate reflection of RRR across geographical units.

Where \({W}_{i,p,c,t}\) indicates the net weight of the imported plastic waste of type \(i\) (referring to one of four waste plastics documented in the harmonised system (HS): PE, PS, PVC and others) from the country \(p\) to the country \(c\) of the year \(t\) ; \({{PI}}_{i,{c},t}\) indicates the per-unit price of the imported plastic waste of type \(i\) by country \(c\) in the year \(t\) ; The upper-case \({C}_{i,c,t}\) denotes the operational costs during the mechanical recycling of plastic waste \(i\) in the importing country \(c\) of the year \(t\) , including costs for labour ( \({{LAB}}_{c,t}\) ), electricity ( \({{ELE}}_{i,c,t}\) ), and rent ( \({{RET}}_{c,t}\) ) in Eq. ( 2 ). \({Q}_{i}\) indicates the physical loss of plastic waste of type \(i\) during mechanical recycling. \({R}_{i,c,t}\) indicates the recycling rate of imported plastic waste of type \(i\) in the country \(c\) of the year \(t\) ; \({{PR}}_{i,c,t}\) indicates the per-unit price of recycled plastic of type \(i\) in the importing country \(c\) of the year \(t\) . The lower-case \({c}_{i,c,t}\) denotes the per-unit operational cost when dividing the \({C}_{i,c,t}\) by \({\sum}_{p}{W}_{i,p,c,t}\) .

The minimum recycling rate of imported plastic waste, enabling an economic break-even point, is referred to as the Required Recycling Rate ( RRR ). To derive a consistent unit price of recycled plastics, we used the trade data for plastics in primary forms (i.e. plastic pellets, flakes, etc.) recorded in the UN Comtrade database from 2013 to 2022 36 , which consists of both virgin and secondary plastics. Within the same database, we also accessed the trade data for plastic waste. Both unit prices for plastic waste and primary plastics are determined by dividing the trade values and the net weights between trading countries. Moreover, given that primary plastic includes a broader range of polymer subcategories than the four plastic wastes (waste PE, PS, PVC and others), we further map waste PE with the primary plastics HDPE and LDPE, using a share factor that varies by country (Supplementary Table  6 ). Similarly, the ‘others’ waste category is mapped to primary plastics PET and PP. The unit price of imported plastic waste and primary plastic are calculated as follows:

Where \({V}_{i,p,c,t}\) and \({W}_{i,p,c,t}\) indicate the trade value and net weight of imported plastic waste of type \(i\) from the country \(p\) to country \(c\) in the year \(t\) ; \({V}_{{ps},p,c,t}\) and \({W}_{{ps},p,c,t}\) (also subscripts for PVC, HDPE, LDPE, PET, and PP) indicate the trade value and net weight of six types of primary plastic exported from country \(c\) to country \(p\) in the year \(t\) , respectively. By grouping HDPE and LDPE as PE, and PET and PP as ‘Others’, with the share factors of \({r}_{{HDPE},c}\) , \({r}_{{LDPE},c}\) , \({r}_{{PET},c}\) , and \({r}_{{PP},c}\) in the country \(c\) , six primary plastics are mapped to four plastic waste types.

Processing bilateral trade data

We analysed 186,861 bilateral trade entries of plastic waste and primary plastics reported from both 22 research countries and their trading partners from 2013 to 2022 in the UN Comtrade database. These entries include trade value and net weight for four plastic waste types (waste PE in Harmonized System (HS) code 391510, waste PS in HS 391520, waste PVC in HS 391530, waste others in HS 391590) and six primary plastic types (HDPE in HS 390120, LDPE in HS 390110, Expandable PS in HS 390311, PVC in HS 390410, PET in HS 390760, PP in HS 390210).

Each trade entry typically includes details such as reporting country, partner country, period, net weight, and trade value. Ideally, each trade flow should be reported by both importer and exporter during the same period, with closely aligned net weight and trade values. However, discrepancies often arise due to varying reporting conventions; exporters typically report trade values as Free On Board (FOB), while importers report them on a Cost for Insurance and Freight (CIF) basis. For our analysis, we require trade values for a country’s imported plastic waste and its exported primary plastics. There are two options: using trade values reported by the research country, including plastic waste import (CIF basis) and primary plastics export (FOB basis), or using mirror trade values reported by the trading partner of the research country, including plastic waste export (FOB basis) and primary plastics import (CIF basis). In calculating the RRR via the cost-benefit equation, we aim to incorporate the international transport cost for importing plastic waste while excluding the transport revenue for recycled primary plastics. Therefore, we prioritise using plastic waste imports (CIF basis) and primary plastics exports (FOB basis), both reported by the research countries. However, for a robustness check, we also include the RRR  results based on mirror trade data reported by the trading partners of the 22 research countries in Supplementary Table  1 .

We detected trade value outliers through a distributional analysis of the value-to-mass ratio for all trade entries by plastic type and year 37 , 38 , 39 . Since most of these unit price distributions follow a lognormal pattern, we transformed them into normal distributions by taking the natural logarithm (ln($/kg)) 40 . By identifying outliers greater than three standard deviations from the mean value, ~2.4% and 2.6% of plastic waste and primary plastic trade entries were flagged as outliers, respectively.

After organising the trade values by research country, year, and plastic type, any empty trade values were replaced with mirror data when available. For example, if the trade values of China’s waste PVC imports in 2022 were missing, the corresponding export values by its trading partners for 2022 were used instead. This replacement accounted for 2.6% and 2.5% of grouped plastic waste and primary plastic entries, respectively. Further details regarding the stepwise changes in trade entries when processing the original trade data are provided in Supplementary Table  7 .

Costs and physical loss

The complete costs from importing plastic waste to producing recycled plastics include plastic waste imports, operational costs (electricity, labour, land rent, water, fuel, transportation, maintenance), fixed asset investments (buildings, machinery, equipment), potential environmental costs, and taxes 18 , 19 , 41 . Based on the work of Uekert et al. 19 , Larrain et al. 18 and Faraca et al. 41 , we consider the four costs consistently across research countries from 2013 to 2022: imports (as indicated by the UN Comtrade database), electricity, labour and rent.

The labour cost for recycling 1 kg of plastic waste was calculated by multiplying the labour input intensity (the required person-hours to recycle one kilogram of plastic waste) by the hourly earnings of employees in each country. The labour input intensity was determined by the recycling company’s annual output and its number of employees, sourced from voluntary disclosures on independent recycling company websites and reports by industry associations, as presented in Supplementary Table  8 . The production rate (expressed in kg/person-hours) was derived by dividing the company’s annual recycling output by the number of its employees and the yearly working hours, which are standardised as 8 h a day and 365 days a year. Subsequently, the labour input intensity was obtained by taking the inverse of the production rate, and these values are averaged at the country level (see Supplementary Table  8 ). The hourly earnings of employees (by manufacturing industry) during 2013–2022 across countries were referenced from the statistics on ‘Average monthly earnings of employees by sex and economic activity (annual)’ by the International Labour Organization 42 , as shown in Supplementary Table  9 .

The electricity cost for recycling per-unit plastic waste is derived by multiplying per-unit electricity consumption varied by plastic wastes and the industrial electricity price across countries and years. We sourced the electricity consumption per plastic waste in mechanical recycling from the life cycle inventories (Ecoinvent and LCA Commons) and other literature (detailed in Supplementary Table  10 ). For European countries, the industrial electricity tariffs were obtained from Eurostat (see Supplementary Table  11 ) 43 . For other countries, the tariffs were gathered from governmental documents or power company announcements, as shown in Supplementary Table  12 .

The rent for recycling 1 kg of plastic waste is based on the area required to recycle 1 kg of plastic waste per year (m 2 /kg*yr), which is calculated by dividing the land area occupied by a recycling company (plants are included) by its annual recycling output (see Supplementary Table  13 ). This value is subsequently multiplied by the annual industrial rent across countries and years, primarily sourced from either yearly or quarterly reports of real estate companies (see Supplementary Table  14 ). The physical losses of four plastic wastes during mechanical recycling stem from prior literature (see Supplementary Table  15 ), where the input of plastic waste and the output of recycled plastic are collected.

Sensitivity and uncertainty analyses

A one-at-a-time sensitivity analysis was conducted to determine how the alteration of key variables impacts the RRR . The six key variables include four costs (imports, labour, electricity, and rent), product price, and physical loss during the mechanical recycling process. Two values (the minimum and maximum value) presenting the pessimistic and optimistic cases for each variable were selected in a 10-year series (2013–2022) across countries.

The probability range of the RRR was studied with a Monte Carlo simulation, where all six variables are included. Referring to Larrain et al. 18 , the product price of recycled plastic is modelled by the normal distribution, with the value of mean and standard deviation evaluated from the 10-year time series across countries. The costs for labour, electricity, rent, and physical loss during the recycling process are assumed to have a Pert distribution, with the minimum value, maximum value, and the most likely value (median value) selected across countries. The import cost is considered to fit a modified Pert distribution with a most likely value weight equivalent to the minimum and maximum value 44 . The resulting uncertainties are propagated with a Monte Carlo simulation (sampling of 30,000) using kernel density smoothing 45 .

Reporting summary

Further information on research design is available in the  Nature Portfolio Reporting Summary linked to this article.

Data availability

The required recycling rate data by country, year, and plastic waste type generated in this study have been deposited under Zenodo https://doi.org/10.5281/zenodo.8328894 , along with other supporting data. In case of questions or requests, please contact K.L.  Source data are provided with this paper.

Code availability

All Python analysis codes and a catalogue have been deposited in Zenodo https://doi.org/10.5281/zenodo.8328894 .

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Acknowledgements

K.L. would like to thank the support from the China Scholarship Council (No. 202006050026). Special appreciation is extended to Yanan Liang at Leiden University for her insightful discussions on the plastic waste trade.

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Li, K., Ward, H., Lin, H.X. et al. Economic viability requires higher recycling rates for imported plastic waste than expected. Nat Commun 15 , 7578 (2024). https://doi.org/10.1038/s41467-024-51923-4

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Advancements and future directions in waste plastics recycling: From mechanical methods to innovative chemical processes

• Chen, Shaoqin
• Hu, Yun Hang

Important progress in plastic waste recycling is discussed. Advantages and limitations of chemical recycling methods are emphasized. Prospects of innovative, sustainable recycling technologies are explored. Perspectives and future research directions of plastics are proposed.

• Plastic Recycling;
• Waste Management;
• Mechanical Recycling;
• Chemical Recycling;
• Circular Economy

Sustainable Construction Exploration: A Review of Multi-Recycling of Concrete Waste

• Review article
• Open access
• Published: 31 August 2024
• Volume 18 , article number  103 , ( 2024 )

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• Jeonghyun Kim   ORCID: orcid.org/0000-0003-2571-5152 1  

This paper provides an overview of literature on the multiple-time recycling of concrete waste and meticulously analyzes the research findings. The paper begins by reviewing the characteristics of recycled materials such as recycled coarse aggregate, recycled fine aggregate, and recycled powder obtained from concrete waste in relation to the recycling cycle. The influence of each of these materials on the mechanical properties and durability of next-generation concrete is analyzed. Moreover, this paper introduces strategies reported in the literature that aim to enhance the performance of multi-recycled concrete. Lastly, this paper identifies and highlights limitations and research gaps, while providing insightful recommendations to drive future exploration of multi-recycling of concrete.

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Literature review on multi-recycling of concrete waste.

Summary of effects of multi-recycling on properties of recycled materials/products.

Review of strengthening methods for multi-recycled concrete.

Identifying research gaps and proposing directions for future studies.

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Studying the usability of recycled aggregate to produce new concrete

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Introduction

The utilization and recycling of concrete waste originated from the need for urban reconstruction following World War II (Buck 1977 ). In the present day, extensive research on recycling concrete waste has been undertaken to address issues such as landfill scarcity, the depletion of natural resources, and increased concern over environmental protection. Although most previous studies report that the performance of concrete containing recycled materials is inferior to that of concrete based on natural materials (Guo et al. 2018 ; Kim et al. 2016 ; Lin et al. 2023 ; Padmini et al. 2009 ), promising results have been reported, demonstrating the ability to produce concrete with performance similar to that of concrete made with natural materials through the application of various methods: reduced material replacement rate (Etxeberria et al. 2007 ; Ozbakkaloglu et al. 2018 ); improved mixing technique (Hiremath and Yaragal 2017 ; Sicakova and Urban 2018 ; Tam et al. 2007 ); material carbonation and CO 2 curing (Tam et al. 2020 ; Zhan et al. 2013 ); enhanced material quality (Kim 2022 ; Wei et al. 2021 ). Furthermore, replacing natural materials with recycled materials in concrete can reduce waste generation (Hossain et al. 2016 ; Martínez-Lage et al. 2020 ) and lower concrete production costs (Suárez Silgado et al. 2018 ). Due to these benefits, many studies have discussed various strategies to encourage the practical utilization of recycled concrete materials (De Brito and Silva 2016 ; Kim 2021 ; Ma et al. 2023 ; Makul et al. 2021 ).

Consequently, actual cases of utilizing recycled aggregates in real-scale concrete structures are being reported worldwide (Kim 2021 ; Li 2008 ; Poon and Chan 2007 ; Silva et al. 2019 ; Xiao et al. 2022a ; Yoda and Shintani 2014 ). Concrete containing recycled materials can generally be referred to as recycled concrete, but the term ‘recycled concrete’ used in literature commonly implies one-time recycling. Tomosawa et al. ( 2005 ) emphasize that if a recycled product cannot be recycled again, it will merely contribute to waste generation for the next generation. Therefore, recycling should aim to reproduce identical products in the original sense of the term, creating a loop.

In this context, there has been a growing interest in the multi-recyclability of concrete waste as a genuine contribution to sustainable development, with studies on this subject being consistently documented (Brito et al. 2006 ; Kim and Jang 2024 , 2022 ). The mechanical properties, durability, and economic and environmental benefits of multi-recycled concrete have been investigated by a few researchers. However, investigations into the repeated recycling of concrete have taken place relatively recently, resulting in a scarcity of comprehensive review on this topic. Therefore, the objective of this study is to conduct a thorough literature review on the multi-recycling of concrete waste. Commencing with an explanation of the conceptual distinctions between one-time and multiple-times recycling of concrete, this study reviews the effects of repeated recycling of concrete on the characteristics of recycled materials obtained from it. This paper also analyzes the effect of using these recycled materials on the properties of next-generation concrete and introduces methods to enhance the performance of multi-recycled concrete. Then, it reviews the environmental and economic analyses of multi-recycled concrete. To conclude, limitations and gaps of the literature are identified, and recommendations for further research are provided.

Multi-Recycling of Concrete

The concept of multi-recycling of concrete waste is presented in Fig.  1 . After undergoing specific recycling technologies, often involving crushing, concrete waste turns into recycled materials. In general, concrete waste can yield recycled coarse aggregate (RCA), recycled fine aggregate (RFA), and recycled powder (RP) (hereafter, recycled materials mentioned in this article refer to RCA, RFA, and RP), and concrete incorporating them can be classified into recycled coarse aggregate concrete (RCAC), recycled fine aggregate concrete (RFAC), and recycled powder concrete (RPC), respectively. This recycled concrete, which has undergone recycling once, is referred to as first-generation recycled concrete (RCAC1, RFAC1 and RCP1 depending on the recycled material used). Crushing the first-generation recycled concrete yields recycled materials again, which have been recycled twice (RCA2, RFA2 and RP2). These are prefixed with ‘multi’ and are designated as multi-recycled coarse aggregate (multi-RCA), multi-recycled fine aggregate (multi-RFA) and multi-recycled powder (multi-RP). And concrete incorporating these multi-recycled materials is termed multi-recycled concrete: multi-recycled coarse aggregate concrete (multi-RCAC), multi-recycled fine aggregate concrete (multi-RFAC), multi-recycled powder concrete (multi-RPC). An explanation of the key terms used in this study is provided in Table  1 .

Concept of multi-recycling of concrete (based on (Kim and Jang 2022 ))

Characteristics of Multi-Recycled Materials

Given that the quality of recycled concrete materials is one of the many factors influencing the properties of concrete (Kim 2022 ), it is of great significance to comprehend the influence of multi-recycling of concrete on material characteristics. Various indicators represent the quality of aggregates, with density and water absorption being the most commonly reported, and occasionally, attached mortar content, Los Angeles abrasion, and crushing index are reported.

The fluctuations of density and water absorption of RCA are shown in Fig.  2 a and b, respectively, clearly indicating a decrease in density and an increase in water absorption for increasing recycling cycles. Over three times of recycling, the density of multi-RCA can be reduced from 4.2% (Lei et al. 2023b ) to up to 24% (Huda and Alam 2014 ) than that of natural coarse aggregate (NCA). For water absorption, NCA exhibits clustered values of 0.32–1.8%, whereas the values for RCA and MRCA vary from 4.45 to 11.2%. In the same recycling cycle, the difference between the minimum and maximum absorption values tends to gradually increase as the number of recycling cycles increased. In previous studies, apart from a study by Yang et al. ( 2022 ), the variation between the minimum and maximum absorption values for RCA1 was 1.95%, the value that escalated to 2.73% and 3.74% for the second and third generations RCAs (RCA2 and RCA3, respectively). In Fig.  2 a and b, the lines representing both density and water absorption exhibit a steep slope during the first recycling (from NCA to RCA1) and then gradually level out during the second recycling (from RCA1 to RCA2) and the third recycling (from RCA2 to RCA3). As noted by Abreu et al. ( 2018 ), the characteristics of coarse aggregate stabilize as the recycling cycle increases.

Characteristics of coarse aggregates over the number of recycling cycles: a density; b water absorption; c LA abrasion; d crushing index (Abed et al. 2020 ; Abreu et al. 2018 ; Chen et al. 2020 ; Huda and Alam 2014 ; Kim and Jang 2022a; Lei et al. 2023b ; Salesa et al. 2022 , 2017a , b ; Visintin et al. 2022 ; Yang et al. 2022 ; Zhu et al. 2016 , 2019b )

Figure  2 c and d show the Los Angeles abrasion and crushing index of coarse aggregates. As the recycling cycle increases, there is a corresponding increase in both the abrasion and crushing index. Higher values for these properties indicate that the aggregate is more susceptible to abrasion and (Mohajerani et al. 2017 ; Zhang et al. 2017 ), suggesting that repeated recycling weakens the aggregate. In contrast, in studies by Salesa et al. ( 2017a , b ), the abrasion resistance of RCA1 was found to be stronger than that of NCA, but the studies did not note whether the RCA1 was obtained from concrete made with the NCA. Nonetheless, when comparing only RCA1, RCA2, and RCA3 used in these studies, it becomes clear that increasing the recycling cycle weakens the abrasion resistance of aggregate.

The above trend is also observed in RFA and RP. Figure  3 a and b show the change in density and water absorption of fine aggregate, and Fig.  3 c shows the change in powder density over the recycling cycle. The fine aggregate in the study by Jung ( 2023 ) and the powder in the study by Kim et al. ( 2023b ) were obtained by crushing multi-RCAC, and the fine aggregate in the studies by Zhu et al. ( 2018 , 2019a ) was obtained from multi-RFAC. However, a common observation across these studies is that as the recycling cycle of concrete increases, density decreases and water absorption increases.

Characteristics of recycled materials over the number of recycling cycles: a density; b water absorption of fine aggregate; c density of cement and recycled powder (Jung 2023 ; Kim et al. 2023b ; Zhu et al. 2018 , 2019a )

Based on the variations in the characteristic over recycling cycles (i.e., decreased density, increased water absorption, abrasion, and crushing index), it can be expected that multi-recycling of concrete diminishes recycled materials quality and is responsible for the poor performance of concrete with those materials. This degradation is attributed to the attached mortar, which makes them looser, more porous, and less rigid than natural materials (Tam et al. 2021 ). Figure  4 shows the attached mortar content in RCA as a function of recycling cycles. Kim et al. ( 2023a ) and Zhu et al. ( 2016 , 2019b ) reported attached mortar content of 32%, 55%, and 62% over three times of recycling, while Thomas et al. ( 2018 ) reported attached mortar content of up to 88% at the given recycling cycles.

Attached mortar content of coarse aggregate over the number of recycling cycles (Chen et al. 2020 ; Kim et al. 2023a ; Thomas et al. 2018 ; Zhu et al. 2016 , 2019b )

The increase in attached mortar content with increasing recycling cycles is associated with changes in the proportion of materials that make up the concrete. Since recycled aggregate contains a certain amount of attached mortar, the volume fraction of recycled concrete is larger than that of natural aggregate concrete (NAC). As a result, as the recycling cycle increases, the fraction of aggregate in concrete decreases, and the fraction of total mortar (fresh mortar and attached mortar) increases. Therefore, more recycled concrete consists of a larger volume of mortar, and the aggregate obtained from it has a higher attached mortar content (Fig.  5 ).

Material proportions of various concretes: a illustration of concretes with natural-, recycled-, and multi-recycled aggregates; b cross-section of the concretes (Thomas et al. 2020 )

Properties of Multi-Recycled Concrete

The attached mortar in RCA, RFA and RP increases their water absorption. Therefore, when water compensation methods, such as adding extra mixing water and increasing the plasticizer dosage, are not applied, the slump of recycled concrete is generally lower than that of natural concrete. Figure  6 shows the slump of RCAC, RFAC and RPC without the water compensation. As the number of recycling increases, the slump decreases noticeably. This can be attributed to the gradual increase in the water absorption of RCA and RFA, as reviewed in the previous section.

Variation in concrete slump over recycling cycles (Huda and Alam 2014 ; Jung 2023 ; Kim et al. 2023a ; Kim and Jang 2022 ; Salesa et al. 2017a )

Hence, to achieve comparable slump values, additional water is demanded, and the quantities of additional water reported in previous studies are listed in Table  2 . As expected, as the replacement rate and the number of recycling cycles increase, higher quantities of water are required. When replacing 25% of NCA with RCA, it demands 5.2% to 6.9% more mixing water during three times of recycling cycles, whereas 100% replacement requires 28.9% more water (Abreu et al. 2018 ). Similarly, RFA requires 6.4%, 19.2%, and 25.2% of additional water at replacement rates of 25%, 75%, and 100% to achieve similar slumps in the third recycling cycle (Zhu et al. 2018 ).

Due to the presence of additional factors influencing slump, such as particle shape and the moisture state of recycled materials, which were not addressed in the original research articles, an intercomparison between studies was not performed.

Air Content

An adequate level of air content in concrete improves its frost resistance (Hosseinzadeh and Suraneni 2021 ; Tanesi and Meininger 2007 ), while both insufficient and excessive air content can cause mechanical properties and durability-related issues (Özcan and Emin Koç, 2018 ; Wang et al. 2022a , b ). Hence, some specifications specify permissible air content ranges for concrete under specific exposure conditions (e.g. 3.5–7.5% depending on aggregate size for ASTM C94 (ASTM C94/C94M-21b, 2021 )).

Typically, recycled material-based concrete exhibits higher air content compared to natural material-based concrete. This is attributed to factors such as rough surface textures, greater angularity, and the presence of pores in the attached mortar (Silva et al. 2018 ). As the multi-recycling increases the attached mortar content, the air content in concrete increases progressively in proportion to the recycling cycle (Fig.  7 ). When concrete is recycled multiple times as coarse aggregate (i.e., multi-RCAC), the air content increases gradually. Huda and Alam ( 2014 ) reported air content of 3.6%, 3.9%, and 4.4% for the 1st, 2nd, and 3rd generations, respectively. The air content of NAC was 3.4%. Similar results were also reported in the following literature (Yang et al. 2022 ). Salesa et al. ( 2017a ) even reported no change in air content in the 1st and 2nd generations. Considering the tolerance of air content (e.g. ± 1.5% for ASTM C94 (ASTM C94/C94M-21b 2021 )), the effect of repeated recycling on the air content of concrete can be acceptable. However, unlike RCA, RFA can significantly affect the air content (Silva et al. 2018 ). In a study conducted by Jung (Jung 2023 ), the air content of concrete containing 30% RFA during the three generations was 5.4%, 6.2%, and 7.3%, showing a sharp increase compared to NAC (4.3%).

Variation in concrete air content over recycling cycles (Huda and Alam 2014 ; Jung 2023 ; Salesa et al. 2017a ; Yang et al. 2022 )

Compressive Strength

The most fundamental property of hardened concrete is its compressive strength. Figure  8 a and b show the 28-day compressive strength for multi-RCAC in absolute and relative scales, respectively. Most previous studies agree that multi-recycling has an unfavorable effect on the compressive strength of concrete. The compressive strength of RCAC decreases to 82.1–96.4% of NAC in the first recycling cycle, 83.4–93.8% in the second recycling cycle, and 57.6–90.1% in the third recycling cycle. The strength loss in RCAC is a consequence of the increased content of attached mortar in RCAs as the number of recycling cycles increases. As discussed earlier, RCAs become more porous with an increasing number of recycling cycles. Additionally, the compressive strength decreases with each recycling cycle due to a variety of complex factors, including the instability of the interfacial transition zone (ITZ) and the formation of micropores and cracks in RCA resulting from repeated crushing processes (Abreu et al. 2018 ; Huda and Alam 2014 ; Lee and Choi 2013 ; Zhu et al. 2019b ).

Compressive strength of multi-recycled coarse aggregate concrete over recycling cycles in absolute ( a ) and relative scales ( b ) (Abreu et al. 2018 ; Huda and Alam 2014 ; Lei et al. 2023b ; Salesa et al. 2017a ; Visintin et al. 2022 ; Yang et al. 2022 ; Zhu et al. 2016 , 2019b )

Conflicting trends have been found in the following studies (Salesa et al. 2017a ; Visintin et al. 2022 ). Salesa et al. reported an increase in compressive strength of 4.4–5.1% over three recycling cycles compared to that of NAC. The authors concluded that high-quality RCA obtained from precast members and the presence of unhydrated cement in the RCA contributed to the improvement in compressive strength. A similar case was also observed in the study by Kim et al. ( 2023a ). In that study, in which precast concrete members were crushed and used as RCA in concrete repeatedly, the compressive strength of RCAC up to the second recycling cycles was 99–111% of that of NAC. In a study by Visintin et al. ( 2022 ), the compressive strengths of RCAC1, RCAC2 and RCAC3 were 3.1–9.8% higher than that of the control concrete. The authors noted that the internal curing by the additional mixing water to compensate for the high water absorption of RCA would have resulted in the similar compressive strengths over the three times of recycling. According to a study by Domingo-Cabo et al. ( 2009 ), when the effective water-cement ratio is constant, several properties of concrete (slump, compressive strength and elastic modulus) can be similar regardless of RCA replacement rate, and Eckert and Oliveira ( 2017 ) reported that extra mixing water can improve the ITZ structure without significantly affecting the effective water-cement ratio. However, other studies, where the effective water-cement ratio was kept constant for concretes by adjusting additional water, report a decrease in compressive strength with the recycling cycle (Abreu et al. 2018 ; Zhu et al. 2016 , 2019b ). While there may be various factors contributing to these conflicting results, Sosa et al. ( 2021 ) highlight the uncertainty arising from the absence of a reliable method to quantify the actual effective water-cement ratio.

The compressive strength of concrete containing RFA and RP is shown in Fig.  9 . The change in compressive strength with respect to the recycling cycle is consistent with that of RCAC, i.e., a decrease in compressive strength as the recycling cycle increases. In particular, RP can cause significant strength loss at relatively low replacement rates, which is attributed not only to the micropores and cracks in RP itself but also to the replacement of cement by RP, reducing hydration products (Kim et al. 2023b ; Kourounis et al. 2007 ).

Compressive strength of multi-recycled fine and powder concretes over recycling cycles in absolute ( a ) and relative scales ( b ) (Kim and Jang 2022 ; Zhu et al. 2018 , 2019a )

Based on the above review, it can be concluded that, in general, multi-recycling has an unfavorable effect on the compressive strength of concrete. Identifying the factors that contribute to the deterioration of properties in repeatedly recycled concrete is essential for sustainability. Practically, it is nearly impossible to track how many times concrete has been recycled. Although studies specifically designed to investigate the effects of multi-recycling may use 100% recycled aggregate, it is uncommon for recycled aggregate to entirely replace natural aggregate in real-world structures. Additionally, industrial regulations in some countries restrict high replacement rates (Tam et al. 2018 ). Due to this complexity, it is necessary to identify the factors that lead to the deterioration of properties in multi-recycled concrete. To understand the relationship between the characteristics of recycled aggregates and the properties of the concrete containing them, Fig.  10 illustrates how the density and water absorption of recycled aggregates correlate with the compressive strength of the concrete, irrespective of the number of recycling cycles. Generally, an increase in aggregate density enhances the compressive strength of the concrete, whereas a higher aggregate water absorption diminishes it.

Relationship between specific gravity of aggregates and compressive strength of concrete ( a ) and water absorption of aggregates and compressive strength of concrete ( b ) (Yang et al. 2022 ; Salesa et al. 2017a ; Huda and Alam 2014 ; Abreu et al. 2018 ; Lei et al. 2023b ; Kim et al. 2023a ; Kim and Jang 2022 ; Zhu et al. 2016 , 2018 , 2019a , b )

Tensile Strength

Tensile strength is one of the crucial mechanical property of concrete since concrete cracks tend to occur in tension, exerting a significant influence on crack formation under load (Zain et al. 2002 ). Figure  11 illustrates the variation in tensile strength over recycling cycles. Generally, with an increase in the number of recycling cycles, the tensile strength decreases. This trend is commonly observed irrespective of the type of recycled materials. On rare occasions, some studies have reported an increase in tensile strength with an increase in recycling cycles. For instance, in a study by Huda and Alam ( 2014 ), the tensile strength of RCAC1 and RCAC2 was observed to be 3–4% higher than that of NAC. The authors interpreted this phenomenon as a decrease in the water-cement ratio in the ITZ due to absorption of mixing water by the RCA. Consequently, the reduced water-cement ratio enhances the bond between RCA and the cement paste. However, in the case of R3, the tensile strength sharply decreased, and the authors attributed this to the low quality of RCA and the multiple layers of ITZ. Similar results were also reported by Yang et al. ( 2022 ).

Tensile strength of multi-recycled concretes over recycling cycles: a and b concrete with multi-recycled coarse aggregate in absolute and relative scale; c and d concrete with multi-recycled fine aggregate and powder in absolute and relative scale (Yang et al. 2022 ; Visintin et al. 2022 ; Huda and Alam 2014 ; Abreu et al. 2018 ; Kim et al. 2023a ; Kim and Jang 2022 ; Zhu et al. 2016 , 2018 , 2019b )

Drying Shrinkage

Drying shrinkage occurs when water in the pores of the cementitious matrix evaporates in a dry environment (Wu et al. 2017 ). Due to the characteristics of recycled materials, such as low stiffness, high porosity, and water absorption, recycled concrete has weak resistance to shrinkage deformation (Mao et al. 2021 ; Wang et al. 2020 ; Wu et al. 2022 ; Xiao et al. 2022b ). The characteristics of recycled materials further deteriorate with repeated recycling, causing concrete recycled for more cycles to exhibit greater shrinkage than concrete recycled for fewer cycles. Silva et al. ( 2021 ) and Kim et al. ( 2023a ) recorded the drying shrinkage of multi-RCAC for 91 days, respectively. Silva et al. ( 2021 ) reported that drying shrinkage is associated with an increase in both aggregate replacement rates and recycling cycles (Fig.  12 a). Similarly, Kim et al. ( 2023a ) noted that drying shrinkage increases as the recycling cycle increases but suggested that the mix design method, so called an equivalent mortar volume (EMV) method (Fathifazl et al. 2009 ; Kim et al. 2016 ; Yang and Lee 2017 ), which deducts the amount of new mortar equal to the amount of mortar attached to RCA, can help suppress drying shrinkage. According to the study, the drying shrinkage of EMV-based concrete with 100% RCA in the 1st-, 2nd- and 3rd generations was 8.5%, 12.2%, and 5.5% lower than that of concrete proportioned by a traditional mix design (Fig.  12 b).

Drying shrinkage of multi-recycled aggregate concrete by Silva et al. ( 2021 ) ( a ) and Kim et al. ( 2023a ) ( b )

One notable difference between the two studies is shrinkage deformation at early ages: Silva et al. ( 2021 ) found that the shrinkage behavior of NAC and RCAC was similar regardless of the recycling cycle up to 7 days, whereas the study by Kim et al. ( 2023a ) showed clear differences in drying shrinkage deformation caused by the recycling cycle at 7 days. In the former study, the moisture compensation for achieving consistent workability was carried out at each recycling cycle, while in the latter case, it was not. Additional water absorbed into RCA is later released, acting as a moisture source for curing, and this internal curing effect can delay the initial drying shrinkage of multi-recycled concrete (Yildirim et al. 2015 ; Zhang et al. 2013 ; Zhutovsky and Kovler 2017 ).

Water Absorption

Water is the main transport medium for the penetration of harmful substances such as chlorides and sulfides into the pore structure of concrete (Wang et al. 2019 ). Therefore, understanding the movement of water in concrete is important from a durability perspective and some researchers have investigated the relevant properties. Figure  13 shows water absorption of concrete by immersion. As expected, the absorption capacity increases with the number of recycling cycles, and this trend is similarly observed in absorption through capillary action as shown in Fig.  14 . Both of these properties are associated with porosity (Silva et al. 2021 ). Due to the presence of attached mortar, which increases with repeated recycling, recycled aggregates and RP exhibit higher porosity and water absorption capacity than natural aggregates and cement, respectively, leading to more permeable pores. These changes affect the absorption capacity of the next-generation concrete (Salesa et al. 2017b ).

Water absorption of multi-recycled concretes over recycling cycles (Kim et al. 2023a ; Salesa et al. 2017a , b ; Silva et al. 2021 ; Thomas et al. 2020 ; Visintin et al. 2022 )

Water absorption of multi-recycled concretes over recycling cycles by immersion ( a ) and capillary action ( b ) (Silva et al. 2021 )

Chloride Penetration Resistance

Chloride resistance is a key indicator of concrete durability. Due to variations in the quality of recycled materials, as discussed in previous sections, the resistance of multi-recycled concrete to chloride penetration weakens with an increasing number of recycling cycles. Zhu et al. ( 2019b ) and Kim et al. ( 2023a ) demonstrated a weakening of chloride resistance in RCAC due to repeated recycling, based on the increase in electrical conductivity with recycling cycles. In the former study, the charge passed during three recycling cycles increased from 1537 to 3300 C, while in the latter study, it increased from 2931 to 4331 C over three recycling cycles. Silva et al. ( 2021 ) and Zhu et al. ( 2019b ) also reported an increase in the chloride diffusion coefficient of RCAC by 47.4% and 85%, respectively, compared to that of NAC after three recycling cycles. The deterioration in chloride resistance can be more pronounced when RFA is repeatedly recycled. In another study conducted by Zhu et al. ( 2018 ), the diffusion coefficient of concrete using 100% RFA ranged from 1.33 × 10 –12 m 2 /s to 3.50 × 10 –12 m 2 /s over three recycling cycles, which was 233%, 419%, and 614% higher than that of NAC. Nevertheless, with the chloride diffusion coefficient of multi-RCAC and multi-RFAC satisfying the 100-year design life requirement in severe environments specified in the Chinese code (GB 50010-2010), the authors concluded that the feasibility of multi-recycling of concrete is promising.

Carbonation Resistance

Carbonation is a chemical reaction where hydrated cement paste reacts with CO 2 . This promotes a decrease in the pH of concrete, which is also associated with the corrosion of reinforcement bars.

The lower quality of recycled materials, characterized by low density, high porosity, and microcracks compared to natural materials, is known to promote CO 2 influx, reducing the carbonation resistance of concrete containing them (Silva et al. 2015 ; Tang et al. 2018 ). As multi-recycling further deteriorates these characteristics of recycled materials, a gradual decrease in carbonation resistance of multi-recycled concrete with increasing recycling cycles is expected, and indeed, such experimental results have been reported in studies (Silva et al. 2021 ; Zhu et al. 2019a ).

The carbonation resistance of multi-recycled concrete is further deteriorated in aggressive environments. After undergoing 300 freeze–thaw cycles, the carbonation depth of RCAC3 increased by more than double (117.3%) compared to the concrete before freeze–thaw action (Liu et al. 2021 ). Furthermore, the carbonation depth of RCAC1 and RCAC2, exposed to chloride ions, increased by 2.5 times and 2.7 times, respectively, compared to their pre-exposure levels (Chen et al. 2020 ). Both freeze–thaw action and chloride penetration loosen the pore structure of concrete, increasing its porosity. This increased porosity facilitates the CO 2 diffusion, resulting in a decrease in carbonation resistance. Nonetheless, the authors emphasize the promising result that the carbonation resistance of multi-recycled concretes exposed to harsh environments satisfied the 50-year design service life requirements of the design code (JGJ/T193-2009 and GB/T 50476-2019).

Frost Resistance

Frost resistance of concrete refers to the ability of concrete to withstand freeze–thaw cycles without significant damage and is a key parameter that determines the service life of concrete in cold regions. Generally, recycled materials absorb more water and this absorbed water is discharged into the cement paste, weakening its cold resistance. Zhu et al. ( 2019b ) investigated the frost resistance of multi-RCAC. During 800 freeze–thaw cycles, both the dynamic elastic modulus and weight decreased in the order of NAC, RCAC1, RCAC2, and RCAC3 (i.e., NAC has the highest modulus and weight, while RCAC3 has the lowest). In particular, RCAC3 after 600 cycles showed a higher mass loss than RCAC2 after 800 cycles of freeze–thaw, clearly indicating a deterioration in frost resistance as concrete was repeatedly recycled. To complement this, Wang et al. ( 2022a , b ) have stated that, in order to maintain a multi-cycle recycling system in an environment subject to freeze–thaw action, the parent concrete needs to be a high-performance concrete to prevent durability damage during its service life.

Microstructural Analysis

The scanning electron microscope results of concretes undergoing three cycles of recycling are shown in Fig.  15 . For NAC, one ITZ between the NCA and the fresh mortar is observed, along with a few microcracks due to moisture evaporation (Fig.  15 a). As recycling progresses multiple times, the cement matrix becomes complex. In RCAC1, there are two ITZs: ITZ1 between NCA and the existing hardened mortar, and ITZ2 between this RCA1 and the new mortar (Fig.  15 b). RCAC2 has three ITZs, including the two observed in RCAC1 and ITZ3 between RCA2 and the fresh mortar (Fig.  15 c). RCAC3 shows four ITZs, including the three observed in RCAC2 and ITZ4 between RCA3 and the fresh mortar (Fig.  15 d) (Belabbas et al. 2024 ). The ITZ is a weak point where concrete is more prone to cracking. In particular, the ITZ between new and old mortar provides a weaker bond than the ITZ between aggregate and mortar (Zuo et al. 2020 ). This explains why the performance of concrete recycled more times is lower than that of concrete recycled fewer times.

Scanning electron microscopy analysis of concrete with various recycling cycles (Belabbas et al. 2024 )

Performance Enhancement of Multi-Recycled Concrete

Reduction in replacement rate of recycled materials.

One of the simplest way to mitigate performance loss in recycled concrete is to reduce the replacement of natural materials with recycled ones (Bai et al. 2020 ; Kim et al. 2022 ). Some studies have attempted to compensate for the performance loss from multi-recycling by including natural aggregate in each recycling cycle (Abed et al. 2020 ; Marie and Quiasrawi 2012 ; Shmlls et al. 2022 ). For example, in a study by Marie and Quiasrawi ( 2012 ), RCAC1 was prepared with 20% RCA replacement rate (i.e., 80% of the coarse aggregate in RCAC1 was natural aggregate), from which RCA2 was obtained. RCAC2 was prepared with 20% RCA2 (i.e., 80% of the coarse aggregate in RCAC2 was natural aggregate) (Fig.  16 ). As shown in Table  3 , this approach enhanced the workability, mechanical strength, and water absorption of the second generation RCAC. However, it should be noted that the environmental benefits diminish as natural aggregate is used for each recycling cycle. Furthermore, due to the 80% NCA used in RCAC1, the RCA2 obtained from RCAC1 does not truly represent ‘multi-recycled’ aggregate.

Multi-recycling of concrete with and without natural aggregates

Carbonation of Recycled Materials

In recent times, numerous studies have emerged focusing on the utilization of CO 2 in concrete. When the hydration products in RCA, RFA, and RP are exposed to CO 2 , calcium carbonate and silica gel are formed, and this reaction fills the pores and cracks of the recycled materials, making the microstructure dense (Fang et al. 2021 ; Lu et al. 2018 ; Luo et al. 2018 ; Xuan et al. 2017 ). Liu et al. ( 2022 ) applied this carbonation technique to second-generation RCA and investigated the effect of its use on the properties of concrete. The RCA2 was carbonated under the following conditions: a temperature of 20 °C; relative humidity of 55%; CO 2 concentration of 20%, and a CO 2 gas pressure of 0.5 MPa. Table 4 shows the characteristics of RCAs before and after carbonation treatment, and the properties of concrete containing the RCAs. The carbonated RCA2 has better characteristics (higher density, lower water absorption) as an aggregate for concrete than non-carbonated RCA2. The quality of aggregates plays a crucial role in concrete performance (Kim 2022 ); consequently, concrete containing carbonated RCA2 exhibits higher compressive strength, a denser pore structure, and improved durability. In particular, it is worth noting that concrete containing carbonated RCA2 performed better than RCAC1, suggesting the possibility that carbonation treatment can offset performance losses by multi-recycling. A positive effect of carbonated RCA can also be found in other study (Wang et al. 2022a , b ).

Vibration Mixing

Yang et al. ( 2022 ) investigated the effect of vibration mixing on multi-recycled concrete. Table 5 summarizes the properties of vibrated and non-vibrated concretes, showing that the vibrated concrete exhibits better workability and strengths. This performance improvement is attributed to vibration breaking the viscous connection between cement particles, preventing the cement agglomeration and allowing RCA to be better coated with fresh mortar (Xiong et al. 2019 ; Zhao et al. 2021 ). This vibration mixing method has the advantage of being applicable without changing the mixing components of concrete.

As represented in Fig.  5 , due to the presence of attached mortar, recycled concrete exhibits a larger volume of mortar compared to NAC, with a smaller proportion of original aggregates. To control this imbalance in material proportions, the EMV mix design method has been proposed (Fathifazl et al. 2009 ). The primary principle of the EMV method is to offset the volume of fresh mortar by the volume of attached mortar, thereby making the total mortar volume of recycled concrete equivalent to that of NAC. The performance efficiency and environmental benefits of this method have been reported in various literature (Fathifazl et al. 2011 ; Jiménez et al. 2014 ; Rajhans et al. 2019 ; Yang and Lee 2017 ). Kim et al. ( 2023a ) applied the EMV method to multi-cycle recycling. The EMV-based concretes with the same material volume were prepared and tested over three recycling cycles. and the test results are summarized in Table  6 . While concrete designed using conventional methods demonstrated a gradual loss of performance with increasing recycling cycles, the EMV-based concrete exhibited no obvious loss in the performance at each recycling cycle, indicating the importance of mix design that takes into account the characteristics of recycled aggregate.

Use of Plasticizer

As mentioned in the previous section, it was discussed that one of the consequences of multi-recycling is a reduction in concrete workability. In response, Kim et al. ( 2023c ) aimed to improve the workability of RCAC3 by increasing the plasticizer dosage and investigated its influence on the concrete properties. In the study, plasticizer dosages in RCAC3 were increased from 0.8 to 1.2% of cement in 0.1% increments. Except for the plasticizer dosage, the rest of the mix design remained the same, and the control group was RCAC1 with 0.8% plasticizer. Table 7 summarizes the experimental results. The slump of RCAC3 increased with increasing plasticizer dosage. In addition, for the hardened properties, the density, mechanical strength, and capillary absorption were improved, and some properties (tensile strength and capillary absorption) achieved similar performance to that of the RCAC1 as the plasticizer dosage was increased to 1.0%. This is related to the fact that free water due to the increase in plasticizer is used to promote hydration of the cement (Zhao et al. 2021 ). However, this positive effect diminishes when the plasticizer dosage exceeds the threshold. Therefore, the authors recommended determining the appropriate dosage.

Environmental and Economic Analysis of Multi-Recycled Concrete

The environmental aspects of multi-recycling of concrete have been discussed in some studies. In a study by Visintin et al. ( 2022 ), it was found that the benefits of using RCA were insignificant as the process of recycling concrete waste into aggregate is similar to the process of converting natural stone into aggregate. However, the authors noted that the effect of transportation distance should be investigated. Generally, in urban areas, the generation of construction waste and the demand for concrete coexist, resulting in shorter transportation distances for RCA compared to NCA. In the life cycle assessment by Lei et al. ( 2024 ), RCAC1, RCAC2, RCAC3 demonstrated superior environmental performance compared to NAC in terms of Global Warming Potential (GWP), Photochemical Ozone Creation Potential (POCP), Acidification Potential (AP), Eutrophication Potential (EP), and Cumulative Energy Consumption (CED). For example, the GWP of RCACs with three different recycling cycles was 8.5%, 12.1%, and 15.8% lower than that of NAC. The authors attributed these environmental benefits to two main factors: (i) the shorter transportation distance of RCA from construction waste recycling plants to concrete production facilities (20 km) compared to NCA from quarries to concrete production facilities (380 km); (ii) avoiding landfilling of concrete waste through recycling. They particularly emphasized that these effects are amplified when concrete is repeatedly recycled. Similar results were reported in other study (Shmlls et al. 2023 ), where the use of RCAC1 and RCAC2 resulted in approximately 20% and 25% reduction in GWP. Kim and Jang ( 2022 ) analyzed the environmental impact of using multi-RP as a partial replacement for cement. According to their study, incorporating 20% RP recycled three times reduced GWP by 15% while achieving the target strength. Furthermore, it was observed that concrete containing 10% RP offered greater environmental benefits compared to concrete containing 100% RCA, which is due to the significantly higher CO 2 emissions from cement compared to other materials used in concrete.

The economic viability of multi-recycling of concrete has been relatively underexplored, and its benefits can vary depending on the circumstances. In the study by Kim and Jang ( 2022 ), the production cost of RCAC2 was approximately 5% lower than that of NAC. However, the compressive strength was 13% lower, leading the authors to emphasize that economic discussions should consider both intended properties and production costs together. The modification of multi-RAC for better performance could potentially worsen its economic viability. In a study (Shmlls et al. 2023 ), replacing NCA with RCA1 by 30% increased the cost per cubic meter of concrete from $147.4 to $152.9. Even with a 70% replacement rate of RCA1, the cost remained higher than that of NAC at $151.9. Similarly, replacing RCA1 with RCA2 resulted in the costs of $149.0 and $148.1 at 30% and 70% replacement rates, respectively, still higher than that of NAC. This is because natural aggregate is not inherently expensive material, and additionally, more additives are required proportionally to the number of recycling to enhance the workability of multi-RAC. While the primary purpose of recycling is environmentally driven, economic feasibility is essential for sustained implementation in actual industries. Further comprehensive research is necessary to understand the economic viability of multi-recycling.

According to the findings of this review, multiple-time recycling is responsible for the performance loss in concrete, and the extent of this loss becomes more pronounced with an increase in the number of recycling cycles. Nevertheless, it is crucial to comprehend that the loss does not imply restricting the utilization of multi-recycled concrete. Even when no strengthening methods are applied, multi-recycled concrete can be used as normal strength (e.g., 20 MPa) as well as high strength concrete (e.g., 50 MPa) (see Figs.  9 and 10 ), and this strength range satisfactorily meets the requirements in many industry standards. For instance, Zhu et al. (Zhu et al. 2019b ) reported that the 1st and 2nd generation concrete met the 100-year design life requirement in harsh and cold environments according to the Chinese design code, while the 3rd generation concrete satisfied the 50-year design life requirement.

Despite the limited number of publications on multi recycling, several research gaps could be identified. In most previous studies, concrete waste was recycled multiple times as coarse aggregate, and few studies treated it as fine aggregate and powder (Fig.  17 a). Considering that fine particles are inevitably generated during the concrete multiple recycling process (Salesa et al. 2017b ; Zhu et al. 2019b ), further research on their utilization is needed to achieve multi-and-zero waste recycling. It is well known that the properties of concrete are significantly influenced by the materials used. Modern concrete may incorporate additives such as fibers, nanomaterials, water reducers and air-entraining agents to achieve optimal performance (Kidalova et al. 2012 ; Kowalik and Ubysz 2021 ; Sánchez-Pantoja et al. 2023 ). Furthermore, for environmental benefits, industrial by-products like fly ash and ground granulated blast-furnace slag, as well as industrial waste such as glass, brick, clay, plastic, and ceramic wastes, are used as supplementary cementitious materials (Shirdam et al. 2019 ; Sičáková et al. 2017 ; Tawfik et al. 2024 ). However, the effect of the repeated recycling of concrete containing these materials on the properties of next-generation concrete has not been investigated. Another research gap can also be clearly found in the types of concrete. The majority of previous studies have focused on conventional concrete, which requires compaction. Little study was carried out on self-compacting concrete and mortar (Fig.  17 b). Additionally, very limited research has been conducted on steel and fiber reinforced concrete.

Number of studies categorized by material type ( a ) and mixture type ( b )

Table 8 summarizes the types of tests conducted in literature. Workability and compressive strength tests, as the most representative properties of fresh and hardened concrete, were the most frequently performed. Following these, properties such as tensile strength, water absorption and elastic modulus were also often measured. On the other hand, relatively little testing has been performed on the durability of concrete, such as abrasion resistance and freeze–thaw resistance.

In summary, current studies on multi- recycling of concrete primarily focus on the basic properties of traditional compacted concrete containing RCA, which may provide directions for further research. For example:

The utilization of fine particles generated from multi-cycle recycling needs to be studied. The use of fine particles is directly related to achieving zero waste, and their counterparts, i.e., sand and cement, emit more CO 2 than coarse aggregates. Therefore, fine particles have the potential to make a significant contribution to reducing CO 2 emissions.

A more systematic investigation is needed into how the raw materials of parent concrete affect the properties of next-generation concrete. This will help identify which factors have favorable or unfavorable influences on repeated recycling.

Examining various types of cementitious mixtures, including self-compacting mortar and concrete, can provide an expanded understanding of multi- recycling. Furthermore, given that concrete is often used in combination with fibers and rebar, it is essential to investigate the effect of multi-recycling on reinforced concrete.

One notable weakness of recycled concrete is its low durability, which is a major impediment to using recycled materials in concrete. Thus, it is crucial to conduct various tests on the durability properties of multi-recycled cementitious mixtures and explore ways to enhance their performance.

The environmental benefits of multi-recycling of concrete need to be better understood, and an investigation into establishing an economic model to sustain this recycling practice is also necessary.

These exemplary further studies are expected to build the body of knowledge on multi- recycling of concrete and contribute to better utilization of waste.

Conclusions

This paper has conducted a literature review on the multi- recycling of construction waste, and the following conclusions can be drawn:

The number of times concrete is recycled affects the quality of the recycled material obtained from it. As the number of recycling increases, the recycled aggregate and powder have more micro cracks and pores.

Recycled materials downgraded by multi-recycling have a negative influence on the workability, mechanical properties, and durability of concrete.

Performance losses resulting from multiple cycles of concrete recycling can be offset by various strengthening methods, such as carbonation of recycled materials, modified mix design and mixing techniques.

It should be noted that the lack of available data limits a clear assessment of the effect of multi-recycling of concrete and the identification of key contributing factors for the effect. Nevertheless, in this study, clear research gaps in existing studies were identified, and the limitations and potential of multi-recycling of concrete were discussed. Further comprehensive research is needed on the various types and properties of multi-recycled concrete.

Data Availability

The datasets generated during and/or analysed during the current study are available from the corresponding author on reasonable request.

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This research was funded in whole by the National Science Centre, Poland (Grant number 2022/45/N/ST8/01782).

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Kim, J. Sustainable Construction Exploration: A Review of Multi-Recycling of Concrete Waste. Int J Environ Res 18 , 103 (2024). https://doi.org/10.1007/s41742-024-00652-z

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At least 63% of imported plastic waste needed to be recycled for economic viability in 22 top importing countries from 2013 to 2022, exceeding the average domestic recycling rate of 23%.

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Economic viability requires higher recycling rates for imported plastic waste than expected - Nature Communications

Kai and colleagues found that at least 63% of imported plastic waste needed to be recycled for economic viability in 22 top importing countries during 2013–2022, exceeding the average domestic recycling rate of 23%.

Key Factor in Environmental Impacts

There’s a deeper story behind this work. Initially, it wasn’t intended to be part of our research plan. However, we found ourselves at a crossroads when trying to determine the fate of imported plastic waste. This all started by exploring the environmental impacts of plastic waste trade, a topic that’s particularly intriguing given its controversial nature. On one hand, some advocate for the trade as a component of the circular economy, arguing that plastic waste serves as feedstock for secondary production in Global South countries. On the other hand, critics see it as "waste colonialism," believing that imported plastic waste exacerbates environmental impacts in these regions, where mismanagement like open dumping and burning is more likely to occur.

After conducting life cycle assessment, we realized that the environmental consequences of plastic waste trade largely depend on the recycling rate of the imported waste. If more of the imported waste is recycled, there are more avoided impacts from reduced virgin plastic production, leading to less share for waste mismanagement and lower overall environmental impacts. Conversely, if less of the waste is recycled, it results in more mismanagement and increased pollution.

Can domestic plastic recycling rates serve as a reliable proxy?

As an alternative, we considered using domestic plastic recycling rates as an alternative. The logic seemed sound—if a country’s domestic recycling rate is as low as 15%, it’s unlikely that the recycling rate for its imported plastic waste would be significantly higher. We encountered this reasoning across various sources. For exmaple, the Guardian article discussed how the UK exported large amounts of plastic waste to Turkey, suggesting that most of this waste was likely mismanaged, given Turkey’s domestic recycling rate of around 15%, with further evidence of mismanagement from Greenpeace . We  also came across research published in Nature Communications that used domestic recycling rates to evaluate the environmental impacts of plastic waste trade. However, we had a sense that comparing a country's general recycling rate with its recycling rate for imported waste is like comparing apples to oranges.

Two reasons against the proxy of domestic plastic recycling rates

We identified two key reasons why using the domestic recycling rate as a proxy wasn’t appropriate (detailed in our paper). First, the domestic recycling rate is usually calculated based on all domestically generated plastic waste, which is often mixed and   dirty   . In contrast, imported plastic waste is classified as a commodity under the Harmonized System (HS) code, indicating it’s relatively clean and intended for recycling. Second, according to the UN Comtrade database, this "commodity" is purchased by importing countries, incentivizing them to recycle and profit from it, since other waste treatment practices like incineration, landfilling, open dumping, or burning don’t yield any monetary return.

Deducing the Required Recycling Rate (RRR)

A new method focusing on the economic viability of importers came to mind: calculating a required recycling rate (RRR) based on the break-even point between the costs of importing and recycling (including labour, electricity, and real estate rentals) and the value of the recycled plastics.

We assume that the value of recycled plastics is comparable to primary plastics and account for physical losses throughout the recycling process. The import costs and primary plastic values are derived from 186,861 bilateral trade records for four plastic wastes (PE, PS, PVC, and others) and six primary plastics (HDPE, LDPE, PS, PVC, PET, and PP) from the UN Comtrade database. Data on other costs and physical losses are collected by country and year.

Divergent impacts of recycling rates on environmental assessment

The previously metioned paper in Nature Communications  assessed the environmental impact of China's plastic import ban using domestic recycling rates for imported plastics. It estimated the net carbon emissions of treating traded plastic waste in 2018 at 0.13 million tonnes CO2-equivalent (CO2-eq). However, in a scenario reflecting a 50% increase in countries’ recycling rates, closely aligning with our calculated RRR for the same period, this figure dropped to -60 kilotonnes CO2-eq. These varying environmental impacts highlight the significant difference between using RRRs and domestic recycling rates to assess the environmental consequences of global plastic waste trade.

Acknowledgments

We extend our gratitude to everyone who contributed to and sponsored this work.  Dr. Hauke Ward , who assisted in refining the methodology and in every creative discussion. Prof. Hai Xiang Lin , who suggested that a higher recycling rate might be achieved if the importer covers the cost of plastic waste.  Prof. Arnold Tukker , who first questioned the misuse of a country's general plastic recycling rate and encouraged me to seek better data in July 2023.  Special appreciation is extended to Dr. Yanan Liang , whose discussions reminded me of which side actually bears the cost of the waste trade, providing crucial perspective.

We also appreciate the support from the Institute of Environmental Sciences (CML) at Leiden University for sponsoring the publication of this work. 

If you’re interested in the full details, feel free to dive into the affiliated paper . For a deeper understanding of the environmental impacts calculated using the RRR method, check out our other paper published in Environmental Science & Technology .

We sincerely hope you find this research insightful and helpful!

All the figures in this post are produced by generative AI to aid understanding.

I am focused on evaluating the environmental impacts of the global plastic waste trade. My work involves reconciling bilateral trade data, evaluating recycling rates, modeling environmental consequences, and analyzing the effects of related policies. Beyond these technical aspects, I am also deeply interested in understanding the broader societal and environmental impacts of trade and climate change policies.

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Solar photovoltaic module end-of-life waste management regulations: international practices and implications for the kingdom of saudi arabia.

1. Introduction

1.1. literature survey, 1.2. research methodology, 1.2.1. research design, 1.2.2. data collection, 1.2.3. analytical framework, 1.2.4. criteria for analysis.

• Comprehensiveness: the scope and depth of regulations covering the life cycle of solar PV modules.
• Stakeholder Engagement: the involvement of manufacturers, consumers, and recyclers in the EOL process.
• Enforcement and Compliance: mechanisms in place to enforce regulations and ensure compliance.

1.2.5. Link to Objectives and Research Questions

• What are the best practices in solar PV EOL waste management among leading countries?
• How can these practices be adapted to fit the Saudi context under Vision 2030?
• What are the potential benefits of implementing these international practices in Saudi Arabia?

1.2.6. Inclusion and Exclusion Criteria

• The policy document must address solar photovoltaic (PV) end-of-life waste management, with a focus on recycling, reuse, and disposal of PV components.
• The policy document must be accessible to the public and written in English.
• The policy document must have been published by 2024.
• The policy document does not pertain to solar PV end-of-life waste management.
• The policy document is not publicly accessible or is written in a language other than English.
• Legal Framework : evaluation of the legal and regulatory structures governing solar PV end-of-life waste management in each country.
• Policy Goals : analysis of the objectives outlined in the policies concerning solar PV end-of-life waste management.
• Policy Tools : review of the instruments utilized to achieve the goals related to solar PV end-of-life waste management.
• Implementation : examination of the effectiveness of policy implementation regarding solar PV end-of-life waste management.
• Policy Effectiveness : assessment of how successful the policies have been in managing solar PV end-of-life waste.
• Stakeholder Engagement : Analysis of the involvement of key stakeholders, including industry and civil society, in the development and execution of solar PV end-of-life waste management policies.

2. Solar PV Module Waste Composition

Composition of solar pv waste, 3. countries generating higher solar pv end-of-life waste volumes, 4.1. national solid waste law.

• The new Solid Waste Law stipulates that “solid waste” refers to items and substances that are in a solid, semi-solid, or gaseous state, contained in containers, and are generated from various activities such as production, daily life, and other activities. These items and substances have lost their original usefulness, are discarded or abandoned, despite potentially still having value.
• In the new Solid Waste Law, “solid waste” refers to objects and substances that are subject to management under laws and administrative regulations, except for waste that has undergone treatment to reduce its volume and hazardousness, meets national product quality standards, and does not pose a risk to public health or the environment. Additionally, any items that do not meet the standards and procedures for solid waste identification are not classified as solid waste.
• Producers’ Responsibility : the law requires producers of products to establish a sound EOL management system and to bear the primary responsibility for the collection, transportation, and disposal of EOL waste generated by their products.
• Collection and Disposal : Producers of products are required to set up collection points for EOL waste generated by their products and to ensure the proper disposal of such waste. In addition, producers are required to publish information about the EOL waste management system on their websites and in product manuals.
• Environmental Protection : Producers are responsible for ensuring that the disposal of EOL and it does not cause harm to the environment or human health. This includes proper handling and disposal of hazardous materials, such as lead, cadmium, and other toxic substances, which may be present in the panels.
• Reporting and Record-Keeping : producers of products are required to submit annual reports on the EOL waste management activities and to maintain records of the EOL waste collected and disposed of.
• Public Information : the law requires producers of products to provide information to the public about their EOL waste management activities and to promote public awareness of the importance of proper EOL waste management.
• Penalties : Companies that violate the provisions of the National Solid Waste Law, including those related to the management of EOL solid waste, may face fines and other penalties. The amount of the fine will depend on the severity of the violation and the extent of any environmental damage caused.

4.2. Specifications for Recycling and Reusing Thin-Film Solar Panels in Construction Applications (GB/T 38785-2020)

• Strategies for the collection, transit, and processing of discarded panels, extraction, and refinement of valuable components, alongside secure management and elimination of toxic substances;
• Establishes benchmarks for assessing the environmental repercussions of recycling and repurposing thin-film solar panels, alongside recommendations for the architectural and manufacturing phases to enhance recyclability and reusability;
• Seeks to advance the photovoltaic sector’s ecofriendly growth by advocating for the conscientious disposal of waste panels and effective resource utilization.

4.3. Regulations for the Control of Pollution from Storage and Landfill of Nonhazardous Industrial Solid Waste (GB 18599-2020)

• Comprehensive guidelines for the design, construction, operation, and closure of nonhazardous industrial solid waste storage and landfill sites;
• Criteria for site selection, groundwork, sludge management, gasses emission control, and overall site supervision to avoid pollution;
• Directions on acceptable waste categories for these sites, detailing procedures for waste reception, processing, and transportation;
• Requirements for continuous monitoring, documentation, and emergency response plans to efficiently identify and address environmental risks;
• Emphasizes community engagement in waste management processes and outlines methods for effective communication with local communities and stakeholders.

4.4. Technical Guidelines for the Recycling of Electrical and Electronic Equipment Waste (GB/T 23685-2009)

• Detailed procedures for the gathering, storing, transport, and dealing of WEEE, including the design and operational standards for collection facilities and safe handling practices for hazardous components;
• Guidelines for the extraction and purification of valued materials from WEEE, like metals and plastics. It contains techniques for material separation, reprocessing processes, and the potential reutilization of components;
• Standards for assessing the environmental effect of WEEE recycling activities, considering energy consumption and greenhouse gas emissions, to promote sustainability in recycling practices;
• The importance of community awareness and education regarding the proper management of WEEE, including recommendations for educational campaigns and community engagement efforts;
• Encouragement of the advancement of a healthy market for recycled materials to support a circular economy and decrease the waste generation.

4.5. Definitions Related to the Recuperation of Waste Products (GB/T 20861-2007)

• Offers a detailed compilation of definitions and terms relevant to the recovery of waste materials, encompassing aspects like sorting, collection, transportation, and processing of waste;
• Covers a wide range of terms related to the recovery and recycling operations for various waste materials, including plastics, metals, paper, and glass;
• Introduces vocabulary related to the environmental and economic advantages of waste recovery, emphasizing efficient resource use, principles for carbon footprint assessment, and circular economy;
• Objectives for standard harmonization to facilitate waste recovery global collaboration and communication;
• Highlights the reputation of exact and uniform terminology to enhance mutual cooperation and understanding between shareholders in waste recycling and recovery industries.

4.6. Observations

5. the united states of america (usa), 5.1. national legislation: resource conservation and recovery act (rcra).

• Solid Waste Definition Clarification: A new guideline was established to delineate when discarded PV panels are considered as solid waste, exempting them from RCRA mandates under certain conditions. This clarification aids the solar sector in navigating EOL waste management with reduced regulatory ambiguity;
• Hazardous Waste Regulation Adjustments: a conclusive regulation now omits specific PV panel types from being treated under hazardous waste guidelines, alleviating the solar industry’s regulatory obligations and fostering the recycling and repurposing of PV panels;
• PV Panels Conditional Exclusion: a conditional exclusion has been formulated for PV panels managed under specific criteria, enabling their recycling in a manner that is both environmentally responsible and exempt from the RCRA’s stringent regulations;
• Recycling Standards for PV Panels: the EPA has set forth standards for the recycling processes of PV panels, ensuring their environmentally sound management and recycling practices;
• National Recycling Capacity Assessment for PV Panels: An analysis to gauge the United States’ recycling industry’s capacity to handle PV panel waste was conducted. This assessment is instrumental in shaping future policy directions for PV panel waste management.

5.2. Overview of State-Level Regulations for Managing End-of-Life Waste from Solar PV Systems in the U.S.

5.3. observation, 6.1. waste management and public cleansing law (1970), 6.2. the resource recycling act (2013), 6.3. promotion of recycling of small waste electrical and electronic equipment (small appliance recycling act) (2013), 6.4. japan photovoltaic energy association (jpea) recycling guidelines (2014).

• Some of the key recommendations in the guidelines include the following:
• Manufacturers and importers of PV panels should establish a system for the collection and disposal of their products at the end of their useful life;
• Recycling companies should be certified by the government and follow appropriate safety and environmental regulations;
• PV panels should be dismantled and recycled to the extent possible, with materials such as glass, aluminum, and copper separated and sent for recycling;
• Hazardous materials contained within PV panels, such as lead and cadmium, should be managed and disposed of properly.

6.5. Ministry of the Environment’s Guidelines for the Sound Material-Cycle Society (2018)

6.6. observations, 7.1. the national programme on solar pv waste management provides a framework for managing eol solar pv waste (2020).

• Developing a comprehensive regulatory framework for the managing of PV EOL;
• Establishing a mechanism for the collection, transportation, and storage of PV EOL waste;
• Creating a system for the environmentally sound disposal of EOL solar PV waste;
• Promoting research and development in the area of solar PV waste management;
• Building capacity for the management of EOL solar PV waste;
• Creating awareness among stakeholders about the importance of sustainable solar PV waste management.

7.2. The E-Waste (Management) Rules (2016)

7.3. cpcb guidelines on the environmentally sustainable management of eol solar pv waste (2018), 7.4. observations, 8.1. the electrical and electronic equipment act (elektrog) (2005), 8.2. the waste electrical and electronic equipment directive (weee) (2012), 8.3. the german solar association (bsw), 8.4. observations, 9. global concern about solar pv end-of-life waste recycling and management, 9.1. challenges in recycling solar pv waste, 9.1.1. technological and economic barriers, 9.1.2. regulatory and logistical issues, 9.2. potential environmental impacts, 10. the kingdom of saudi arabia, 10.1. saudi arabia waste management law, 10.2. observations, 11. lesson learned for the ksa.

• Extended Producer Responsibility (EPR) : mandating that manufacturers take responsibility for the entire life cycle of their products, including the design, take-back programs, and covering recycling costs.
• Public Awareness and Education Campaigns : launching initiatives to educate the public about the importance of recycling and proper waste management practices.
• Public–Private Partnerships (PPPs) : fostering collaborations between the government and private sector to develop and operate recycling facilities and waste management programs.
• Extended Producer Responsibility (EPR) : implement policies that require manufacturers to manage the life cycle of their products, ensuring they are recyclable and facilitating take-back programs.
• Public Awareness and Education Campaigns : develop and execute campaigns to raise public awareness about EOL waste and encourage community participation in recycling efforts.
• Public–Private Partnerships (PPPs) : encourage partnerships to invest in and manage recycling infrastructure and waste management programs effectively.
• Development of Recycling Infrastructure : invest in specialized facilities and technologies to handle the anticipated increase in solar PV waste.

12. Conclusions and Policy Implications

• Detailed policy analysis to support the development of robust EOL waste management regulations.
• Exploration of advanced technological solutions for recycling and disposal.
• Strategies for effective stakeholder engagement and public awareness.
• Economic assessment of EOL waste management practices to ensure sustainability.

Author Contributions

Data availability statement, acknowledgments, conflicts of interest.

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Ali, A.; Islam, M.T.; Rehman, S.; Qadir, S.A.; Shahid, M.; Khan, M.W.; Zahir, M.H.; Islam, A.; Khalid, M. Solar Photovoltaic Module End-of-Life Waste Management Regulations: International Practices and Implications for the Kingdom of Saudi Arabia. Sustainability 2024 , 16 , 7215. https://doi.org/10.3390/su16167215

Ali A, Islam MT, Rehman S, Qadir SA, Shahid M, Khan MW, Zahir MH, Islam A, Khalid M. Solar Photovoltaic Module End-of-Life Waste Management Regulations: International Practices and Implications for the Kingdom of Saudi Arabia. Sustainability . 2024; 16(16):7215. https://doi.org/10.3390/su16167215

Ali, Amjad, Md Tasbirul Islam, Shafiqur Rehman, Sikandar Abdul Qadir, Muhammad Shahid, Muhammad Waseem Khan, Md. Hasan Zahir, Asif Islam, and Muhammad Khalid. 2024. "Solar Photovoltaic Module End-of-Life Waste Management Regulations: International Practices and Implications for the Kingdom of Saudi Arabia" Sustainability 16, no. 16: 7215. https://doi.org/10.3390/su16167215

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Chemical plastics recycling is ready to go: Researchers show that it's all about the stirring

by Fabio Bergamin, ETH Zurich

Hundreds of millions of tons of plastic waste are generated worldwide every year. Scientists are working tirelessly on new methods to recycle a large proportion of this waste into high-quality products and thus enable a genuine circular economy. However, current recycling practices fall short of this goal.

Most plastic waste is recycled mechanically: shredded and then melted down. Although this process does result in new plastic products, their quality deteriorates with each recycling step.

An alternative to this is chemical recycling, which avoids loss of quality. This method involves breaking down long-chain plastic molecules (polymers) into their fundamental building blocks (monomers), which can be reassembled into new, high-quality plastics, creating a truly sustainable cycle.

Fuels from plastic waste

As the approach of chemical recycling develops, the initial focus is on breaking down these long polymer chains into shorter-chain molecules that can be used as liquid fuels, say, or lubricants.

This gives plastic waste a second life as petrol, jet fuel or engine oil. Scientists at ETH Zurich have now laid down important foundations for developing this process. These enable the global scientific community to engage in more targeted and effective recycling development work.

Researchers in a group led by Javier Pérez-Ramírez, professor of catalysis engineering, investigated how to break down polyethylene and polypropylene with hydrogen. Here, too, the first step is to melt the plastic in a steel tank. Gaseous hydrogen is then introduced into the molten plastic. The work has been published in Nature Chemical Engineering .

A crucial step involves adding a powdered catalyst containing metals such as ruthenium. By carefully selecting a suitable catalyst, researchers can increase the efficiency of the chemical reaction, promoting the formation of molecules with specific chain lengths while minimizing byproducts such as methane or propane.

Rotational speed and geometry are key

"The molten plastic is a thousand times thicker than honey. The key is how you stir it in the tank to ensure the catalyst powder and hydrogen get mixed right through," explains Antonio José Martín, a scientist in Pérez-Ramírez's group.

Through experiments and computer simulations , the research team demonstrated that the plastic is best stirred using an impeller with blades parallel to the axis. Compared to a propeller with angled blades or a turbine-shaped stirrer, this results in more even mixing and fewer flow vortices. The stirring speed is equally crucial. It must be neither too slow nor too fast; the ideal speed is close to 1,000 revolutions per minute.

The researchers successfully developed a mathematical formula to describe the entire chemical recycling process with all its parameters. "It's every chemical engineer's dream to have a formula like this at hand for their process," Pérez-Ramírez says. All scientists in the research field can now precisely calculate the effect of the stirrer's geometry and speed.

With this formula, future experiments can focus on directly comparing different catalysts with the influence of mixing under control. In addition, the principles developed here are central for scaling up the technology from the laboratory to large recycling plants.

"But for now, our focus remains on researching better catalysts for the chemical recycling of plastics," Martín says.

Journal information: Nature Chemical Engineering

Provided by ETH Zurich

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Researchers Turn to Cottonseed Oil to Replace Toxic Chemicals in Textiles, Paper Products

Cotton is one of the world’s most important crops. Its white, fluffy fibers have been spun and woven into nearly every type of textile — from shirts and pants to towels and sheets — for centuries. Now its seeds could be key to replacing the toxic chemicals used to produce wrinkle-free fabrics and water-resistant paper products. 

With a $294,000 grant from the U.S. Department of Agriculture, NC State professors Richard Venditti and Sunkyu Park are exploring the potential of using cottonseed oil to create non-allergenic, biodegradable products that can be used as finishing agents for cotton apparel and wet-strength agents for paper. 

“By developing these bio-based textile finishing agents and wet strength additives, we hope to provide more environmentally sustainable alternatives to petroleum-based products used in the textile and paper industries,” said Venditti, the Elis-Signe Olsson Professor of Pulp and Paper Science and Engineering at the College of Natural Resources. 

Finishing agents are chemicals applied to fabrics to reduce or eliminate wrinkles and to improve the quality and durability of fabrics, while wet-strength agents are chemicals applied to paper towels, toilet paper and other paper products to improve their resistance to moisture and other fluids.

Textile manufacturers often use finishing agents that contain formaldehyde to produce wrinkle-resistant fabrics. Formaldehyde, at certain levels, can have harmful effects on plants, animals and humans. 

Paper manufacturers use synthetic resins that contain polyaminoamide-epichlorohydrin or glyoxalated polyacrylamide to produce water-resistant filter paper, toilet paper and paper towels. When paper is recycled or flushed, the chemicals can enter into wastewater streams and negatively impact the environment. 

“The introduction of these chemicals and pollutants into the environment is an incredibly difficult challenge for remediation and cleanup efforts and are most likely only solved by replacement of these chemicals with environmentally benign additives,” Venditti said. 

Since 2022, Venditti and Park have worked alongside Cotton Incorporated, a North Carolina-based nonprofit that conducts research and promotion to increase the demand for and profitability of cotton, and HeiQ ChemTex, which develops and manufactures specialty chemicals, to prepare the modified cottonseed oil and test it.  

The researchers, with assistance from Cotton Incorporated and HeiQ ChemTex, have already converted the cottonseed oil into a water-resistant coating for paper products. They’ve also created an emulsion — a mixture of cottonseed oil and water — for application on textiles and paper products.

“We’ve done some preliminary work on fabrics and noticed some changes to properties like texture and softness,” Venditti said. “Right now we’re looking at better defining its wrinkle-resistance, water-resistance and oil-resistance. But it shows a lot of potential.” 

Cottonseed oil has several advantages over soybean oil and other vegetable oils for developing finishing agents and wet strength additives, according to Venditti. It’s not only less likely to cause allergic reactions, unlike soybean and peanut oils, but it’s also less likely to impact global food prices and supplies. 

While cottonseed oil is used by the food service industry, it’s not as prevalent as soybean and canola oils due to its higher price. Venditti and Park hope to eventually commercialize the process of utilizing modified cottonseed oil for finishing agents and wet strength additives to help increase the demand for cottonseed oil.

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