Eco-Design of Energy Production Systems: The Problem of Renewable Energy Capacity Recycling

Abstract

Due to the rapid development of recycling technologies in recent years, more data have appeared in the literature on the environmental impact of the final stages of the life cycle of wind and solar energy. The use of these data in the eco-design of modern power generation systems can help eliminate the mistakes and shortcomings when planning wind and solar power plants and make them more eco-efficient. The aim of this study is to extend current knowledge of the environmental impacts of most common renewables throughout the entire life cycle. It examines recent literature data on life cycle assessments of various technologies for recycling of wind turbines and photovoltaic (PV) panels and develops the recommendations for the eco-design of energy systems based on solar and wind power. The study draws several general conclusions. (i) The contribution of further improvements in PV’s recycling technologies to environmental impacts throughout the entire life cycle is insignificant. Therefore, it is more beneficial to focus further efforts on economic parameters, in particular, on achieving the economic feasibility of recycling small volumes of PV-waste. (ii) For wind power, the issue of transporting bulky components of wind turbines to and from the installation location is critical for improving the eco-design of the entire life cycle.

1. Introduction

The transition to a circular economy can significantly contribute to the achievement of the Sustainable Development Goals (SDGs), in particular, Goal 12 (“ensuring sustainable consumption and production patterns”). The foundations for the successful circulation of resources are laid long before the manufacturing of the products, namely, at the design stage of the production systems. A better design can make a product more durable, more suited for repair, modernization, or restoration. Thus, a better design of manufacturing processes will reduce the environmental burden throughout the entire life cycle of the product.

In the last decade, there has been a growing interest in the issues of optimal choice of energy technologies in academic literature. For instance, Turconi and co-authors [1] compared the main technologies of electricity generation based on hard coal, lignite, natural gas, oil, nuclear power, and several renewable sources by the amount of emissions. Poinssot and colleagues [2] weigh environmental footprints of a closed cycle of nuclear energy against the environmental footprint of an open cycle. Paper [3] examined carbon emissions and water consumption of electricity generation from pulverized coal, wind power, and solar PV, while [4] did the same kind of comparative analysis for a wider mix of technologies, including oil, natural gas, hydropower, biomass, and nuclear energy. Chang and coauthors [5] tried to choose a more ecologically friendly technology from coal and shale-gas-fired power generation, using the amount of greenhouse gas emissions and water consumption as the criteria. Paper [6] compared the environmental impacts of different PV-technologies, while [7] extended the analysis to include wind power and biogas electricity generation for consideration. The interest in this research topic has been determined by the rapid development of renewable energy (RE), which has a lower negative impact on the environment than traditional hydrocarbon energy throughout the entire life cycle [1,8,9,10]. However, the issues of decommissioning and end-of-life (EoL) treatment of renewable energy capacities, such as solar and wind plants, have not been thoroughly studied [11]. Disposal at landfills or incineration at waste processing plants remains the most common EoL treatment technologies for used photovoltaic modules and blades of wind turbines. Various mechanical, thermal, and chemical technologies of photovoltaic (PV) panel recycling, as well as recycling technologies for composite materials used in blades of wind turbines, are just at the beginning of their industrial applications. Therefore, the estimates of their environmental impact are quite different and not widely understood. There is no sufficient evidence for the comparison of available renewable energy technology by environmental impact throughout the entire life cycle, including the EoL stage. These knowledge constraints make the problem of a better eco-design of energy systems difficult to solve.

The significant growth of economic feasibility of solar and wind energy, achieved in recent years, opens up new prospects for even more rapid development of RE around the world, including the countries previously quite skeptical about renewables. This fully applies to Russia, which in recent years has been trying to reduce its lag from leading countries in the development of RE [12]. The hasty introduction of reactive innovations in the field of RE as a necessary response to the lag in technological development cannot be considered as an advantage from an economic point of view. However, it allows taking the experience and knowledge of other countries into account and making better decisions in the design of energy systems from an environmental point of view. In other words, the existing lag in technological development can be turned into a competitive advantage in the eco-efficiency of energy systems.

The environmental impact of the EoL stage is still a challenging area in the field of life cycle assessment. Slowly growing expertise in large-scale decommissioning of RE plants built 20–25 years ago as demonstration projects [13,14] brings new data and allows managers and policy-makers to formulate and solve the problems of optimal eco-design of energy systems. Monitoring of new technologies of EoL treatment of wind and solar plants is critical for effective regulation of the scale and type of RE installations, the method of recycling, and the location of recycling plants. The solution to such problems will contribute to the development of a circular economy.

The aim of this study is to extend current knowledge of the environmental impacts of most common RE technologies throughout the entire life cycle. It examines recent literature data on the life cycle assessment (LCA) of various technologies for the recycling of wind turbines and PV panels and develops recommendations for the eco-design of energy systems based on renewables. The rest of the paper is organized as follows: Section 2 presents the results of the systematic literature review of state-of-art recycling technologies for PV panels and wind turbines. Section 3 describes the features of inventory and life cycle assessments for the case of RE’s recycling. Section 4 investigates the results of life cycle inventory and assessment of RE’s recycling obtained under different assumptions and within different system boundaries. This section also presents general conclusions regarding the possibilities for improvements in the eco-design of the energy technologies under consideration. Section 5 discusses the practical applications of the study for the eco-design of energy systems in Russia. Section 6 debates the main contribution of the study to the existing pool of scientific literature, as well as its limitations and possibilities for future research.

2. Modern Recycling Technologies for PV Panels and Wind Turbines: A Systematic Literature Review

2.1. Ecology Issues of Renewable Energy Sources (RES)-Capacity Recycling

At the end of 2018, the global installed PV capacity reached 486 GW (REN21, 2019; Figure 1). Given that the average life of modern solar panels is 25 years [15], and the average weight of the most common 200 Wp multi-Si PV model is 16.8 kg [16], the world may face the need to dispose of more than 24 million tons of expired multi-Si PV models by 2035. In the future, these volumes will only increase due to the rapid development of solar energy in all countries [17,18]. According to expert forecasts, PV panels will reach 4500 GW of cumulative installed capacity worldwide by the middle of the century [15].

The PV solar panels contain lead (Pb), cadmium (Cd), and many other harmful chemicals. So far, the most common EoL treatment technology for PV models remains their disposal at landfills. [15]. It can be quite dangerous since harmful chemicals can leak into the ground, causing drinking water contamination [20]. Incineration is also used for PV models, as with regular municipal waste. It is crucial to understand that incinerating all kinds of electronic waste, including PV modules, can release toxic heavy metals into the atmosphere [21]. It is also known that some of the materials in PV modules are persistent and accumulative when released. This can cause long-term adverse ecology effects. Incineration also abolishes the opportunity of recovering raw materials. The only advantage of this method is that PV modules do not need to be separated from other commercial or industrial waste [22].

Industrial recycling capacities of used or defective PV models are currently limited and represented only by the Czech company Retina, several factories of First Solar in the United States, Germany, and Malaysia, Toshiba Environmental Solutions [15], and Veolia’s new factory in France. At the legislative level, only the EU has established the rules for the collection and processing of solar panels in Waste of Electrical and Electronic Equipment (WEEE) Directive (Directive 2012/19/EU). Some countries with developed solar energy, such as South Korea, Japan, and the USA, are actively working on the problem of organizing the recycling of solar waste, while others, including China, are only exploring ways to solve this problem [23].

As has been previously reported in the literature, existing industrial technologies for recycling solar panels make it possible to achieve 95–97% recovery of cadmium and tellurium for thin-film solar cells [24], 90% recovery of glass [15,24], and 80% recovery of silicon for multi-Si PV models [25]. They also achieve 67–81% recovery of aluminum and 80% recovery of ethylene-vinyl acetate (EVA), which is used to fix individual PV models into a single panel. The use of recovered materials in the production cycle reduces resource consumption. It can achieve a drastic reduction in waste; however, it is associated with enormous expenditures of energy and/or chemical materials, as well as the emission of significant amounts of harmful substances into the atmosphere. Therefore, according to the data [16], the recycling of all the components of 1-kW multi-Si PV panels by modern industrial thermal and chemical methods emits 6.32 kg SO₂, 23.4 kg NO₂, 13.8 kg CO₂, as well as 0.97 kg of ammonia (NH₃), 2.59 kg of ethylene, and 4.26 kg of methane.

At the end of 2019, the worldwide installed capacity of wind turbines reached 621 GW and 29 GW for onshore and offshore wind energy, respectively [26] (Figure 2). Considering the historical pace of development of wind energy in the world and an average 20-year lifetime of wind turbines, one can expect that in the coming years, up until 2030, the volumes of decommissioning of wind turbines will grow exponentially.

A closer look at the literature on the EoL of wind turbines reveals that the most difficult components for recycling are the blades and the concrete foundation [27,28]. The blades are currently simply landfilled or, in some cases, incinerated at municipal waste plants. As they are made from composite materials, their recycling is very complicated. In cases where it is technologically possible, the recovered materials have much lower quality than virgin materials [29,30]. It does not allow us to reuse them for the same purposes. In addition, the large sizes of the blades, which can reach 80–90 m in length on modern turbines, make it difficult to transport them to the place of recycling.

The landfill of large volumes of wind turbine blades is a severe environmental and economic problem [31]. The mass of the blades of modern wind turbines reaches 12.37–13.41 tons per 1 MW, and for turbines ending their life in the coming years, it is about 8.34 tons per 1 MW [32]. It means that if in 2020, it is necessary to landfill more than 31,000 tons of blades, in 2025, this amount will be already more than 142,000 tons, while in 2030, this number will be more than half a million tons. Landfilling of such large objects will require their preliminary energy-intensive shredding or cutting.

The EoL treatment of offshore wind turbines is an even more significant problem. Although their share in the total volume of wind turbine installations is much lower (Figure 3), they are more susceptible to adverse environmental effects and have a shorter lifetime. Thus, for example, in 2015, the first 10 MW Yttre Stengrund Swedish wind farm was decommissioned after only 15 years of operation [33]. In 2018, Utgrunden I Marine Wind Park (10.5 MW, Sweden) was decommissioned, which had operated for 18 years. In 2013, the first marine park built in the UK by Blyth (4 MW) ceased to operate after 13 years. Due to the lack of a clear demolition program, it is still abandoned in its original location. The Beatrice Demonstration Project (10 MW, UK) met the same fate.

From a logistical point of view, the process of dismantling a marine wind park is much more complicated. Besides, there is currently no clear understanding of whether it is necessary to completely dismantle the foundation of wind turbines or they must be left in the same place [34]. Several methods are being developed in the literature to extend the lifetime of marine parks, which are based on the common idea of their more or less large-scale re-equipment. Some academicians have suggested replacing the engine, gearbox, and turbine blades and reusing the old tower, foundation, and power grid [33]. Others have suggested replacing all wind farm equipment except the foundation, which can last up to 100 years [35]. The implementation of such re-equipment is expected to reduce costs compared with the construction of new offshore wind farms [33]. Still, it does not entirely solve the disposal problem.

From an economic point of view, the extraction of valuable materials from used blades is of more interest. For example, Vestas’ blades are made up of glass fiber, carbon fiber (CF), epoxy resin, and polyurethane [36]. Carbon fiber (CF), in addition to wind energy, is also actively used in aircraft, automotive, shipbuilding, and other industries. Carbon fiber is produced from light fractions of crude oil, propane, and propylene through a multistage chemical treatment that requires significant energy costs. The production of one kg of carbon fiber requires 165 kWh, while the recovery is only 8.8 kWh. According to [37], the energy intensity of the glass fiber recycling is 10 times less than that of its production.

Some European countries, especially the ones the most committed to the concept of developing a circular economy, prohibit the disposal of unprocessed waste with a high organic component at the legislative level. Due to the presence of carbon fiber in the composition of blades material, the organic component in them reaches 30%. It obliges energy companies to look for other ways to treat old blades from decommissioned wind turbines.

Thus, the issues of optimizing the entire life cycle of the most common renewable energy technologies are far from their final solution. The last stages of the life cycle of solar and wind power plants represent the most prospective area for improvements and development of new technologies.

2.2. Modern Technologies of PV Panel Recycling

The modern industrial algorithm for recycling silicon-based photovoltaic panels begins with their disassembly, during which aluminum (frame) and glass parts (coating) are separated. Almost 95% of the glass is recyclable, and all external metal parts can be reused to form new frames for solar modules [38]. The remaining materials are heat-treated at a temperature of 500 °C, during which the encapsulating plastic evaporates, leaving the silicon elements ready for further processing. In modern recycling plants, evaporating plastic is not released into the environment but used as a heat source for further heat treatment (energy recovery technology), which partially reduces adverse environmental effects [15,16,39].

After heat treatment, 80% of the PV modules (by mass) can be reused, while the rest is subject to additional cleaning. Silicon particles contained in cracked and scratched wafers are etched with acid and then melted for reuse to produce new silicon modules, resulting in an 85% level of recovery of the raw silicon material. The reuse of silicon can significantly reduce the adverse environmental effects of silicon photovoltaic panel manufacturing by saving, in addition to silicon sand itself, a significant amount of energy and water used in the production stage of metallurgical silicon. Based on the data of Huang and co-authors [16], it is possible to estimate the energy savings in the manufacturing of 1 kW of a photovoltaic panel due to the reuse of silicon as 76.2 kWh and water savings as 76.2 L. It avoids an average emission of 123.34 kg of CO₂, 3.36 kg of SO₂, and 0.73 kg of NO_x.

The recycling of thin-film photovoltaic panels is more complicated [21,40]. In the first step, the panel is crushed to a particle size of not more than 4–5 mm, which allows the removal of the lamination that holds the internal materials. Unlike silicon panels, the remaining substance consists of both solid and liquid materials, which need to be separated. Liquids pass through a precipitation and dehydration process to allow the recovery of semiconductor materials. Separation of semiconductor metals is carried out in various ways, depending on the technology used in the manufacture of panels; however, an average recovery rate of 95% is achieved. Solids contaminated with so-called interlayer materials are cleaned by vibration. Furthermore, the material undergoes washing, as a result of which there remains clean glass that must be melted and reused [39].

There are some new technologies for recycling photovoltaic silicon modules at the stage of laboratory research: technology for dissolving a laminating film in an organic solvent (tetrahydrofuran, o-dichlorobenzene or toluene), technology for dissolving with additional ultrasonic treatment, hot cutting, pyrolysis, technology for dissolving in nitric acid, and some others [41,42]. For thin-film modules, in addition to organic solvent dissolution and hot cutting technologies, laser exposure, vacuum processing, and flotation technologies are also being developed [43,44].

Previous studies have shown that the recycling of silicon modules becomes feasible at a scale of at least 19,000 tons per year [45]. Only when such volumes are achieved, the cost of processing decreases due to the economies of scale. The economic feasibility of recycling an entire PV panel is much higher than the expediency of recycling PV modules only, since the metal frame and electrical cables are easier to recycle, and the recovered materials (aluminum, copper) are valued higher [46,47]. The economic feasibility of recycling silicon panels substantially depends on the price of new modules, which has fallen significantly in recent years. For example, for Europe, the commercial attractiveness of silicon recycling is also determined by the lack of its mining (Figure 4). Of all the European countries, only Norway has a share of more than 4% in the global silicon production market.

2.3. Modern Technologies of Blade Recycling

The simplest method of recycling a composite material is mechanical (cutting, crushing, crumpling). However, it damages individual fibers, which leads to a decrease in mechanical characteristics. According to the data [49], the tensile strength of recycled fiberglass compared to new is not more than 78%; for carbon fiber, it is less than 50%, with the fiber recycling ratio (recyclate yield rate) of 55–58% of the initial mass. The material recycled in this way has a low cost and cannot be a substitute for the virgin material. Typically, such a recycled composite material is used for less demanding mechanical applications, for instance, aggregates for artificial wood or asphalt and concrete in the construction industry [50,51]. Recycled carbon fiber (CF) is also suitable for short-fiber nonwoven composites used in aircraft and vehicle interiors [52]. Manufacturers can use recycled material as a filler or strength enhancer in new products, reducing the cost and environmental impact of waste disposal. In addition, recycled CF can be mixed with a polymer to produce thermally and electrically conductive materials used to improve the durability of paint, cement, and other construction materials [36,53].

Other actively developed methods for processing composite materials are pyrolysis, oxidation in a fluidized bed, and chemical decomposition of polymer resins (binders) [54]. During pyrolysis, the composite is heated to 450–700 °C in the absence of oxygen. The polymer resin is evaporated, and the fibers remain intact and can be restored entirely [36]. The recyclate yield rate of both carbon and glass fibers during pyrolysis is 67–70% of the original mass. The residual tensile strength is 52% for fiberglass and 78% for carbon fiber [49].

During oxidation in a fluidized bed, the polymer component (resin) is burned in a stream of hot (450–550 °C) air enriched with oxygen. The carbon fiber recyclate yield rate is higher and reaches 86–90% of the initial mass, while for fiberglass, it is 42–44% [49]. The residual tensile strength is 50% for fiberglass and 75% for carbon fiber. The chemical influence on the binder polymer element, selected directly for a specific combination of fiber and polymer matrix, helps to get the maximum possible amount of fiber suitable for reuse with minimal time and resources. The recyclate yield rate of both carbon and fiberglass is almost 100% [49], the residual tensile strength for fiberglass is 58%, and for carbon fiber, it is 95%.

To the best of our knowledge, the economic characteristics of the various processes of recycling of composite materials are absent in the literature because all of the methods mentioned above are still at the stage of laboratory research. Meanwhile, the data on the current energy intensity of these methods are known, which in combination with data on recycling efficiency, allow us to judge the competitiveness of methods from an economic point of view indirectly (

Table 1. Comparison of end-of-life (EoL) treatment methods for wind turbine blades. Source: based on data of [49].

Method	Fibrous Recyclate Yield Rate, %	Retained Tensile Strength of Recycled Fibre Compared to Virgin Fibre, %	Energy Demand, MJ/kg
Landfill	0	0	0.26
Mechanical	55–58%	GF—78% CF < 50%	0.27
Pyrolysis	67–70%	GF—52% CF—78%	21.2
Fluidised-bed process	GF—42–44% CF—86–90%	GF—50% CF—75%	GF—22.2 CF—9.0
Chemical	100%	GF—58% CF—95%	19.2

). However, economic feasibility depends not only on the cost of the production process but on the availability and size of the market for recycled materials. Therefore, the prospects of the above methods for recycling blades of wind turbines are also determined by the fact that they are suitable for the recycling of composite wastes from other sectors of the economy (aircraft, automotive). Therefore, the cost of their development and practical application can be reduced due to economies of scale.

The possibility of incinerating composite materials in municipal waste plants for energy recovery depends on the ratio of polymer to carbon fiber in its composition [50]. A significant disadvantage of incinerating composite material is that approximately 60% of the initial mass remains in the form of ash. Ashes must also be disposed of in some way. At the moment, it is either subject to disposal at landfills (which also contradicts the legislation in the field of waste management in some countries) or to reuse in construction materials [55].

Another possibility for recycling wind turbine blades is to reuse them as load-bearing structures in the construction of buildings, technical infrastructure (e.g., bridges), or to create artificial reefs [55,56]. However, cases of such reuse of the blades are more likely experimental than industrial methods of disposal.

As can be seen from the analysis of the currently existing technologies for processing used capacities for wind and solar energy, the final stage of the life cycle of renewable energy facilities can impact the environment significantly. Therefore, the choice of electricity generation technology in planning the development of the energy system in a specific territory (region, municipality) should take into account the total environmental effects throughout the life cycle.

3. Methodology and Data

In the last decades, life cycle assessment (LCA) methodology is commonly used in the literature for evaluating the environmental performance of the entire life cycle of the products, processes, or production systems. The LCA algorithm is described in the series of ISO 14040–14043 standards and consists of the following four main stages:

(1) Goal and scope definition. At this stage, a researcher has to determine the research objective, define the functional unit, and set the system’s boundaries. When determining the system’s boundaries, most researchers use the “cradle-to-grave” or “cradle-to-gate” approaches. In a “cradle-to-gate” approach, life cycle stages such as extraction of raw materials, transportation, processing of materials, and manufacturing of the product are taken into account. When using the “cradle-to-grave” approach, steps such as the usage and final disposal of the product are additionally taken into account.

(2) Life cycle inventory (LCI). At this stage, a complete map of the studied life cycle is constructed as a sequence of production and transportation processes. The inputs of each production/transportation process (such as raw materials, water, and energy consumption), intermediate processes, and outputs (such as main product, by-product, wastes, and emissions to the air, water, and soil) are determined. As a rule, a specific production process that occurs at a particular enterprise, including its supply chain, is investigated. If some stage of the life cycle exists so far only as a laboratory (experimental) process and has various scenarios, then averaged data from various sources can be used for calculations. If the processes are not yet profoundly understood, their inventories are incomplete. In such cases, the results of inventory analysis often complement sensitivity analysis.

(3) Life cycle impact assessment (LCIA). At this stage, the identified potential environmental impacts of the production system are translated into such categories as global warming, acidification, ecotoxicity (human, marine, or terrestrial), ozone depletion, abiotic depletion, and eutrophication. Indicators for measuring these environmental impact categories are determined by one of the following mature methods of LCIA: CML 2001, Cumulative Energy Demand (CED), eco-indicator 99, EDIP, ILCD, ReCiPe, IPCC, or IMPACT 2002+. Some methods (for example, CML 2001) allow the translation of all the adverse environmental effects of the product life cycle into physical units, such as kg CO₂-eq, kg NO_x-eq, kg PO₄-eq, and kg 1.4-DCB-eq. Others (for example, ReCiPe endpoint) allow the aggregation of all the negative effects in all categories into a single dimensionless quantity). The choice of methodology mainly depends on the preferences of the researcher and the objectives of the study.

(4) Interpretation. At this stage, a researcher summarizes the LCI and LCIA results, identifies critical points of the life cycle with the most considerable adverse effects, and makes recommendations for possible improvements.

If the life cycle of the product also includes the recycling stage, the modeling and calculation of the total adverse environmental effects are complicated by the fact that the recycled product can enter the production cycle again. In this case, it reduces the need for some raw materials. In modeling the processes of recycling, two main approaches are used: the “cut-off approach” and the “end-of-life approach”. When using the first approach, the recycling efforts are economically allocated among the treatment process and all the recovered materials with a positive economic value. In the second approach, the potential benefits from the usage of recycled materials are calculated by awarding credits for the avoided environmental impacts caused by the primary production of replaced products. At the same time, some researchers use a simplified approach called open-loop recycling (OLR), in which further flows of secondary materials are not taken into account [57]. More details about differences in allocation approaches can be found in [38].

When choosing the technology for generating electricity from several possible options by environmental parameters, we are faced with the need to take into account the full life cycle of the electricity plant, including the recycling of massive generating equipment. The first stages of the life cycle of wind and solar energy are relatively well studied using LCA. For example, [6,39,58] compare a broad spectrum of environmental categories of photovoltaic technologies, [59] studies a life cycle of organic photovoltaics, and [60,61,62] compare GHG emissions of several photovoltaic technologies. Other studies [63,64,65] give life cycle assessments of solar panels manufactured in specific locations. Ecological aspects of the disposal of used equipment have also been studied (e.g., [66]). In contrast, accounting for recycling as a primary route for end-of-life treatment is a complex problem [38].

First, as a rule, when conducting an LCA for an electricity plant, 1 kWh of energy produced over the entire life of the equipment is used as a functional unit. Such a choice of a functional unit is convenient because it allows one to compare the results of LCA for different technologies and make a choice in favor of the most environmentally friendly. In LCA of the recycling of used generating equipment, the mass of the equipment itself (PV panel or wind turbine) usually is considered as a functional unit. Therefore, it is impossible to directly aggregate the LCA results of the first and last stages of the life cycle.

The second difficulty is the lack of a universal approach to determining the system’s boundaries for recycling processes. Some researchers include the collection and transportation of used PV panels and wind turbines, while others suggest that recycling is done directly at RES plant location. The third difficulty is that, as noted above, most of the recycling processes of RES capacities are at the stage of laboratory research, so the data on the inputs and outputs of these processes are minimal and have great uncertainty.

This study systematizes the LCA results of various methods of EoL treatment of used PV panels and wind turbines, brings them to single units of measurement, and converts them to functional units used to evaluate the earlier stages of the life cycle of electricity plants. The conversion to the functional unit of 1 kWh of all estimations from the literature was carried out under the assumption that the capacity factor of PV plant is 14% [12], solar irradiation in the location of PV plant is 1200 kWh/m²/year [12], the weight of the 200 Wp c-Si module is 16.8 kg [16], the weight of 1 m² of the module is 13.2 kg [67], the weight of 1 m² of the CdTe module is 14.63 kg [67], and the module efficiency is 10.5% [67].

The primary difference of the present research in comparison with similar ones is that we did not restrict ourselves with a certain LCIA method and considered a maximally wide spectrum of impact categories. The purpose of our comparative analysis of LCA results is to identify common conclusions that remain valid even with different approaches, assumptions, and system boundaries and can serve as guidelines for the eco-design of energy systems.

4. Results and Discussion

4.1. PV Panels LCA

To operate with LCA results performed for different system boundaries and functional units, we used data only from the sources where the results were given in physical units (not in dimensionless points). For this reason, the analysis did not include the results of studies where environmental impact assessments were presented in normalized values (e.g., [16,68,69,70,71]). Additionally, because recycling technologies are rapidly progressing, we selected for comparison only the results of the decade.

Table 2. Life cycle assessment (LCA) results for PV panels.

Source	Functional Unit	System’s Boundaries	Life Cycle Impact Assessment Method, Software	Results
Lunardi et al., 2018 [41]	1 kg of c-Si PV waste modules	Manufacturing, installation, operation, recycling (thermal/chemical/mechanical) and reuse of recycled materials in a new circle of manufacturing. Transportation is not included.	ReCiPe endpoint, GaBi	Landfill and incineration: Human health 1.6–1.7 × 10−4 DALYs Ecosystems 4.0–4.6 × 10−11 species·year Resources 8–10 $ US Recycling: Human health 0.5–0.6 × 10−4 DALYs Ecosystems 1.1–1.9*10−11 species·year Resources 2.1–4.1 $ US
Latunussa et al., 2016 [72]	1000 kg of c-Si PV waste panels, including internal cables	Delivery of the waste to the recycling plant, recycling (pilot method) PV modules into metallurgical grade silicon scrap, sorting of the recyclable metal and glass fractions, disposal of residues. Transportation within a radius of 500 km by lorry	ILCD midpoint, SimaPro	Abiotic resource depletion (mineral) 4.32 × 10−3 kg Sb eq. CED 2780 MJ Freshwater ecotoxicity 1.31 × 103 CTUe Marine eutrophication 1.05 kg N eq. Freshwater eutrophication 0.0456 kg P eq. Terrestrial eutrophication 11.7 molc N eq. Acidification 2.41 molc H+ eq. Photochemical ozone formation 2.86 kg NMVOC eq Ionizing radiation ecosystem 6.96 × 10−5 CTUe Ionizing radiation human health 22.9 kg U235 eq Particulate matter 0.0819 kg PM2.5 eq Human toxicity (non-cancer) 1.84 × 10−5 CTUh Human toxicity (cancer) 2.83 × 10−5 CTUh Ozone depletion 2.35 × 10−5 kg CFC-11 eq. Climate change 370 kg CO2 eq.
Ardente et al., 2019 [11]	1000 kg of c-Si PV waste panel with TPT (Tedlar-polyethylene terephthalate-Tedlar) back sheet	For innovative FRELP recycling process: Delivery of the waste to the recycling plant, recycling (pilot method) PV modules into metallurgical grade silicon scrap, sorting of the recyclable metal and glass fractions, disposal of residues. Transportation within a radius of 300/500 km by lorry For new pyrolysis recycling process: fluorine-free PV waste is treated in a fixed-bed pyrolysis plant in place of incineration	EDIP2003 and ReCiPe, SimaPro	FRELP/new pyrolysis recycling process Climate change 461.0/361.3 kg CO2 eq Ozone depletion 0.00003/0.00004 kg CFC-11 eq Human toxicity, cancer effects 0.00001/0.00001 CTUh Human toxicity, non-cancer effects 0.00002/0.00001 CTUh Human toxicity (ReCiPe) 349.9/290.7 kg 1,4-DB eq Particulate matter 0.1/0.1 kg PM2.5 eq Ionizing radiation HH 32.0/30.4 kg U235 eq Photochemical ozone formation 3.0/3.8 kg NMVOC eq Acidification 2.7/2.7 molc H+ eq Acidification (EDIP) 52.1/32.5 m2 Terrestrial eutrophication 12.2/12.0 molc N eq Freshwater eutrophication 0.054/0.1 kg P eq Marine eutrophication 1.095/1.1 kg N eq Freshwater ecotoxicity 255.4/246.3 CTUe Mineral depletion 0.004/0.004 kg Sb eq CED Non-renewable 3074.8/2899.4 MJ
Stolz & Frischknecht, 2016 [67]	1 kg of used 3kw c-Si framed module, mounted on a slanted roof	Manufacturing, installation, operation, recycling (mechanical) and reuse of recycled materials in a new circle of manufacturing. Transportation on 500 km by lorry	ILCD Midpoint 2011	Particulate matter 0.0375 kg PM2.5 eq Freshwater ecotoxicity 391 CTUe Human ecotoxicity (non-cancer) 4.55 × 10−6 CTUh Human ecotoxicity (cancer) 5.0 × 10−5 CTUh Abiotic resource depletion 6.18 × 10−3 kg Sb eq. Climate change 27.4 kg CO2 eq.
	1 kg of used 3 kW CdTe unframed module, mounted on a slanted roof	Manufacturing, installation, operation, recycling (mechanical, dissolving, precipitation, and dewatering) and reuse of recycled materials in a new circle of manufacturing. Transportation on 678 km by lorry	ILCD Midpoint 2011	Particulate matter 0.0327 kg PM2.5 eq Freshwater ecotoxicity 197 CTUe Human ecotoxicity (non-cancer) 1.32 × 10−6 CTUh Human ecotoxicity (cancer) 2.05 × 10−7 CTUh Abiotic resource depletion 1.74 × 10−3 kg Sb eq. Climate change 6.05 kg CO2 eq.
Held and Ilg, 2011 [73]	1 m2 of CdTe module	Manufacturing, installation, operation, recycling (mechanical, dissolving, precipitation, and dewatering). Transportation of produced and used modules is not included	CML2001	Module/module with a ground-mounted balance of systems Primary Energy 831/1270 MJ Acidification potential 0.223/0.358 kg SO2-eq Eutrophication potential 1.91 × 10−2/2.90 × 10−2 kg PO4-eq Global warming potential 513/861 kg CO2-eq Photochemical Ozone Creation Potential 1.68 × 10−2/3.18 × 10−2 kg Ethene-eq
	1 kWh of electricity	Manufacturing, ground-mounted installation, operation, recycling (mechanical, dissolving, precipitation, and dewatering). Transportation of produced and used modules is not included	CML2001	Solar irradiation in place of installation 1200/1700/1900 kWh/m2/yr Primary Energy (fossil) 0.43/0.31/0.28 MJ Acidification potential 1.23 × 10−4 /8.68 × 10−5/7.77 × 10−5 kg SO2-eq Eutrophication potential 9.95 × 10−6/7.02 × 10−6/6.29 × 10−6 kg PO4-eq Global warming potential 2.95 × 10−2/2.09 × 10−2/1.87 × 10−2 kg CO2-eq Photochemical Ozone Creation Potential 1.09 × 10−5/7.71 × 10−6/6.90 × 10−6 kg Ethene-eq
Ilias et al., 2018 [74]	1 kg of reusable Si	From the collection, transportation (600 km) and recycling of PV waste (thermal treatment of the PV modules and the parallel removal of the organic EVA in a noble gas environment) to final production of the recovered Si	Gabi 6.0	GWP (100 years) 679.84 kg CO2-eq. ODP 5.50 kg R11/CFC-eq. Human Toxic (Cancer) 2.70 CTUh Human Toxic (Non-Cancer) 3.90 CTUh Particulate Matter 0.54 kg PM2.5-eq. Ionising Radiation 0.26 kg U235-eq. Photochemical ozone formation 1.35kg NMVOC–eq. Acidification 5.95 moles of H+eq. Eco Tox 35.65 CTUe Resource depletion, water 406.11 kg Resource depletion, mineral, fossil 1.549 kg Sb eq.
Corcelli et al., 2018 [75]	1 m2 of EoL c-Si PV panel	The thermal treatment of the decommissioned PV panel and the subsequent recycling of the recoverable fractions. High-rate (HR) recovery scenario assumed: the heat produced by the plastics thermal treatment is recovered and then exploited for hot water generation or for heating purposes within the plant where the process takes place. Transportation is not included.	ReCiPe Midpoint (H) v.1.10	Climate change 2.64 × 101 kg CO2 eq Acidification potential 1.11 × 10−1 kg SO2 eq Freshwater eutrophication 5.51 × 10−3 kg P eq Human toxicity 7.74E+00 kg 1,4-DB eq Photochemical ozone formation 6.30 × 10−2 kg NMVOC TEP 2.30 × 10−3 kg 1,4-DB eq WDP 1.88 × 100 m3 MDP 2.17 × 100 kg Fe eq FDP 8.11 × 100 kg oil eq
		High-rate (HR) and low-rate (LR) recovery scenarios with environmental credits/LR scenarios assume that only the aluminum frame and glass are recycled, and the not-recovered fraction of copper, silicon, and ashes are disposed of in a sanitary landfill. Transportation is not included.	ReCiPe Midpoint (H) v.1.10	HR/LR recovery scenario GWP −8.69 × 101/−1.92 × 101 kg CO2 eq TAP −5.79 × 10−1/−2.29 × 10−1 kg SO2 eq FEP −4.39 × 10−2/−1.31 × 10−2 kg P eq HTP −4.18 × 101/−1.73 × 101 kg 1,4-DB eq POFP −2.98 × 10−1/−1.10 × 10−1 kg NMVOC TEP −7.61 × 10−3/−2.32 × 10−3 kg 1,4-DB eq WDP −2.14 × 100/−5.06 × 10−1 m3 MDP −2.04 × 100/−2.39 × 10−1 kg Fe eq FDP −2.06 × 101/−2.64 × 100 kg oil eq

presents the results of a comparative analysis of the photovoltaic plant’s life cycle.

The results of LCA in [41] were calculated according to the ReCiPe endpoint method with GaBi software. The functional unit was 1 kg of silicon-based PV waste modules. The study considered two types of PV modules (multi- and monocrystalline silicon modules) and several end-of-life treatment scenarios: landfill, incineration, and thermal, chemical, and mechanical recycling. In the case of recycling, the system boundaries included manufacturing, installation, operation, recycling, and reuse of recycled materials in a new cycle of manufacturing. One of the notable features of the study was a separate analysis of the transportation phase. In this case, the impacts do not depend on the recycling technology but the waste collection system and distance from it.

The environmental effects of transporting used modules to the place of recycling or disposal were taken into account in two ways: (1) assuming a transportation distance of 50 km; (2) assuming a transportation distance of 100 km (for the case of recycling only). Modeling results showed that transportation for 50 km increases impacts on human health up to 2.1–2.3 × 10⁻⁴ DALY in case of landfill or incineration, and up to 1–1.2 × 10⁻⁴ in case of recycling. Transportation for 100 km for recycling increases the impact on human health up to 1.6–1.7 × 10⁻⁴ DALY, which is comparable to the landfill or incineration indicators on site (without transportation). Similar results were obtained in two other categories of environmental impact. Transportation for 50 km increased ecosystem effects to 5.8–6.2 × 10⁻¹¹ species·year in case of landfill or incineration and 3–3.5 × 10⁻¹¹ species·year in case of recycling. Transportation per 100 km for processing increased the adverse effects up to 4.5–5.5 × 10⁻¹¹ species·year, which offset all the positive effects of recycling compared to landfill or incineration.

Thus, the results of this analysis demonstrate that the recycling plant should be at most 80 km away from a PV plant. Otherwise, landfill and incineration scenarios have lower impacts when considering human health and ecosystems. This is a critical practical conclusion for the eco-design of the energy system regarding the territorial location of PV-plants. Solar plants must be placed compactly in a relatively small radius from the processing plant, or they should be some kind of mobile recycling plant. In addition, it is essential to consider not only the distance but also the method of transportation.

In [67], an LCIA is made for the recycling of c-Si and CdTe PV modules using both the cut-off and end-of-life approaches. Data were collected from several recycling companies in central Europe. The functional unit was 1 kg of used framed c-Si and unframed CdTe PV modules, but the final results were presented for 3 kW modules, mounted on a slanted roof. The life cycle impact assessment competed with the ILCD Midpoint 2011 method, but only six of the most relevant impact categories were taken into consideration.

For the case of c-Si PV modules, the weighted average of multi- and monocrystalline Si PV modules was considered. It was estimated that the used modules would be transported by truck over a total distance of 500 km. The technology of recycling is mechanical. The desirable outputs of the recycling process are the bulk materials of glass cullets, aluminum scraps, and copper scraps. For the case of CdTe PV modules, the average transport distance from the place of installation to the recycling plant was considered as 678 km (data of First Solar’s recycling facility located in Germany). The recycling process includes shredding and milling the used CdTe PV modules in the first step, then removing and dissolving the semiconductor film as the second step. The LCIA results for both types of panels are presented in Table 2. As one can see, CdTe PV modules are superior in environmental performance to c-Si PV modules in all categories of environmental impact. This conclusion can also be used in the eco-design of energy systems.

Note that even though both LCAs of [41,67] were performed throughout the entire life cycle, they still do not provide a complete picture of all the adverse effects of a PV plant. They do not take into account the productivity of a PV plant at the operation stage. More efficient PV plants (both by capacity factor and by energy conversion coefficient) can produce more useful products (electricity) for their life cycle and thereby reduce the need for the production and installation of additional PV capacity. Therefore, it is more appropriate to use 1 kWh of generated electricity as a functional unit for the LCAs of all energy facilities, including solar panels. Recalculation of the results of these two studies into other functional units (1 kWh of electricity), unfortunately, is impossible, since the location of the installation of solar panels is a source of uncertainty in this study. In locations with a high level of solar radiation, the total adverse environmental effects over the entire life cycle can be lower due to the greater volume of produced useful products (electricity).

Latunussa and co-authors [72] apply the LCA methodology to a pilot process of recycling of crystalline-silicon (c-Si) PV panels on the Italian “SASIL S.p.A” company. The functional unit was 1000 kg of PV waste panels, including internal cables. The system boundaries of the LCA included “gate-to-gate” recycling processes, starting from the delivery of the waste to the recycling plant and ending with the sorting of the different recyclable material fractions and the disposal of residues. The transportation of PV waste to the recycling plant was considered, while the decommissioning of the PV plant was not. Transportation was considered under the assumption that the distance from the PV plant to the nearest collection point of electronic waste is no more than 100 km, and the distance between the collection point and the recycling site is 400 km.

The novel process of recycling has a sequence of physical (mechanical and thermal) treatments, followed by acid leaching and electrolysis. The amounts of energy produced by the incineration process (for example, the incineration of the sandwich layer and plastics from cables) are considered as coproducts and their positive impact calculated as a credit (avoided environmental impact). By contrast, the environmental credits derived for potentially substituted primary materials are not included.

The modeling and calculation were implemented with SimaPro software version 8.0. The ILCD midpoint method was used for the life cycle impact assessment (Table 2). The “mineral, fossil, and renewable resource depletion” impact category is replaced with “abiotic depletion, fossil” and “cumulative energy demand (CED)” in order to distinguish the contributions of energy sources from those of nonenergy materials. The “land use” and “water resource depletion” impact categories were not taken into consideration due to their high uncertainty.

The results of this study do not have direct applications for the eco-design of the power system. However, they can be used to calculate the entire life cycle of c-Si PV modules for any method of installing panels and for any location.

Ardente [11] extended the results obtained in [72] by introducing options for the recycling process that depend on the material of the back-sheet. This study also introduced several additional impact categories. The results indicated that there is little potential to reduce the environmental impact of the recycling process in some categories due to the variation of materials for the back-sheet and recycling technologies for the auxiliary parts of the PV panel.

Held and Ilg [73] considered both individual stages and the entire life cycle of thin-film CdTe modules. They used two different functional units: 1 m² of the module and 1 kWh of energy produced. An interesting feature of the study is the comparison of impacts of the module with and without the balance of systems (BoS). It was revealed that the relative contribution of the BoS on the total impact of the PV power plant is around 35% to 45%. Estimates of environmental impacts for 1 kWh of electricity as a functional unit are made under three different assumptions about the amount of solar radiation in the location of the solar power plant. The disadvantage of the study is the lack of accounting for the effects of transporting new modules to the installation site and used modules to the recycling site. As shown above, transportation can have a significant influence on all environmental impact assessments. In addition, in this paper, the lifetime of the PV modules is assumed to be 30 years, which is an overestimation.

The authors of [74] presented a rather unusual approach to life cycle analysis, which the authors defined as “grave-to-cradle”. They suggested that recycled silicon should replace a virgin material in the production of PV panels and considered the process in terms of industrial symbiosis. Recycling is carried out using thermal and chemical methods.

Corcelli and coauthors [75] investigated two c-Si PV panel recycling scenarios: one with a high level of material recovery and another with a low level. Environmental impacts were calculated in two versions: including credits and excluding them. The disadvantage of this study is the lack of accounting for transportation.

Comparing the environmental impacts of recycling with the impacts of all previous stages of the life cycle presented in EcoInvent (

Table 3. Environmental impacts of different life cycle stages of c-Si PV panels (for 1 kWh ¹).

Impact	P+O+R 2	R	P+O
	[41,67,74]	[11,72,75]	EcoInvent
Human heath, DALYs	1.35 × 10−7 [41]–4.59 × 10−7 [41]		1.66 × 10−10
Abiotic resource depletion (mineral), kg Sb eq	1.67 × 10−5 [67]–4.18 × 10−3 [74]	1.08 × 10−8 [11]–1.17 × 10−8 [72]	9.61 × 10−6
CED, MJ		7.51 × 10−3 [72]–8.30 × 10−3 [11]	1.26 × 100
Freshwater ecotoxicity, CTUe	1.06 × 100 [67]	6.65 × 10−4 [11]–5.53 × 10−3 [72]	1.15 × 10−1
Marine eutrophication, kg N eq.		2.84 × 10−6 [72]–2.97 × 10−6 [11]	1.06 × 10−4
Freshwater eutrophication, kg P eq		1.23 × 10−7 [72]–1.13 × 10−6 [75]	6.33 × 10−5
Terrestrial eutrophication, molc N eq.		3.16 × 10−5 [72]–3.29 × 10−5 [11]	1.02 × 10−3
Acidification, molc H+ eq	1.61 × 10−2 [74]	6.51 × 10−6 [72]–7.29 × 10−6 [11]	6.46 × 10−4
Photochemical ozone formation, kg NMVOC eq	3.65 × 10−3 [74]	7.72 × 10−6 [72]–1.29 × 10−5 [75]	3.33 × 10−4
Ionizing radiation human health, kg U235 eq	7.02 × 10−4 [74]	6.18 × 10−5 [72]–8.64 × 10−5 [11]	7.88 × 10−3
Particulate matter, kg PM2.5 eq	1.01 × 10−4 [67]–1.46 × 10−3 [74]	2.21 × 10−7 [72]–2.70 × 10−7 [11]	8.78 × 10−5
Human toxicity (non-cancer), CTUh	1.23 × 10−8 [67]–1.07 × 10−2 [74]	2.70 × 10−11 [11]–5.40 × 10−11 [11]	3.42 × 10−8
Human toxicity (cancer), CTUh	1.35 × 10−9 [67]–7.29 × 10−3 [74]	2.70 × 10−11 [11]–7.64 × 10−11 [72]	2.35 × 10−9
Ozone depletion, kg CFC-11 eq		6.35 × 10−11 [72]–1.08 × 10−10 [11]	7.62 × 10−9
Climate change, kg CO2 eq	7.13 × 10−6 [74]–7.40 × 10−6 [67]	9.76 × 10−4 [11]–5.41 × 10−3 [75]	8.19 × 10−2
Human toxicity, kg 1,4-DB eq	2.08 × 10−2 [74]	7.85 × 10−4 [11]–9.45 × 10−4 [11]	1.72 × 10−1
Acidification potential, kg SO2-eq		2.28 × 10−5 [75]	5.44 × 10−4
Eutrophication potential, kg PO4-eq			2.35 × 10−4
Photochemical Ozone Creation Potential, kg Ethene-eq			3.26 × 10−4
Resource depletion, water, m3	1.10 × 10−3 [74]	3.85 × 10−4 [75]	1.04 × 10−3

¹—The conversion to the functional unit of 1 kWh was carried out on the assumption that the capacity factor is 14%, the weight of the 200 Wp module is 16.8 kg, and the weight of 1 m² of the module is 13.2 kg. ²—P, manufacturing (production) of PV panel with upstream activities; O, operation; R, recycling.

and

Table 4. Environmental impacts of different life cycle stages of CdTe PV panels (for 1 kWh ³).

Impact	P+O+R	P+O (CdTe)
	[67,73]	EcoInvent
Abiotic resource depletion (mineral), kg Sb eq	8.75 × 10−6 [67]	6.12 × 10−6
CED, MJ	2.85 × 10−1 [73]–4.36 × 10−1 [73]	5.99 × 10−1
Freshwater ecotoxicity, CTUe	9.90 × 10−1 [67]	1.50 × 101
Particulate matter, kg PM2.5 eq	1.64 × 10−4 [67]	5.04 × 10−5
Human toxicity (non-cancer), CTUh	6.63 × 10−9 [67]	8.84 × 10−8
Human toxicity (cancer), CTUh	1.03 × 10−9 [67]	7.58 × 10−9
Climate change, kg CO2 eq	3.04 × 10−2 [73]–2.96 × 100 [73]	3.95 × 10−2
Acidification potential, kg SO2-eq	7.66 × 10−5 [73]–1.23 × 10−4 [73]	4.28 × 10−4
Eutrophication potential, kg PO4-eq	6.56 × 10−6 [73]–9.96 × 10−6 [73]	1.90 × 10−4
Photochemical Ozone Creation Potential, kg Ethene-eq	5.77 × 10−6 [73]–1.09 × 10−5 [73]	1.55 × 10−5

³—Recalculation into a functional unit of 1 kWh was carried out by assuming that the mass of 1 m² of the module is 14.63 kg (according to First Solar), module efficiency 10.5%, solar irradiation 1200 kWh/m²/year.

), one can conclude that any technology for recycling PV waste is more ecologically friendly than landfilling. However, this is true only if the recycling plant is located in the same region where the panels are manufactured and used. In this case, the transportation of heavy modules over long distances was not required. This is consistent with what has been found in [46,47] for economical and in [76] for environmental parameters of recycling. By comparing the results from Table 3 and Table 4, we can conclude that CdTe panels are preferable over silicon panels for the full life cycle (with EoL stage) in most categories of environmental impacts. These results go beyond previous reports [77], showing that current techniques used in the recycling of PVs produce higher impacts in the case of c-Si than in the case of CdTe. A further novel finding is the following: despite the fact that current technologies for recycling of PVs can be significantly improved from an environmental point of view [78], the contribution of these improvements to the negative impacts throughout the life cycle is insignificant. Therefore, it is advisable to focus on further improvement in economic parameters, in particular, on achieving the economic feasibility of recycling small volumes of PV waste.

4.2. LCA of Wind Turbines or Their Components

Aggregating the LCA results for wind turbines, as in the previous case, we also did not consider the results presented in dimensionless units (e.g., [68,79,80]), or received more than ten years ago (e.g., [81]). For a more detailed understanding of the development possibilities of recycling, we also separately examined the studies analyzing the LCA of the composite materials of blades (

Table 5. LCA data for wind turbines and components.

Source	Object/Functional Unit	System’s Boundaries	Life Cycle Impact Assessment Method, Software	Results
Garrett and Rønde, 2012 [82]	Wind park/1 kWh of electricity	Extraction raw materials, manufacturing turbine components and wind plant components, installation, maintenance and end-of-life disposal. Blades are not recyclable.	CML, GaBi DfX	Abiotic depletion (elements) 0.44 mg Sb eq Abiotic depletion (fossils) 0.10 MJ Acidification 37 mg SO2-eq Eutrophication 3.7 mg PO4-eq Freshwater aquatic ecotoxicity 100 mg DCB-e Global warming 7.7 g CO2-eq Human toxicity 1150 mg DCB-eq Marine aquatic ecotoxicity 1100 g DCB-eq Photochemical oxidation 4.1 mg Ethene Terrestrial ecotoxicity 19 mg DCB-eq
Bonou et al., 2016 [83]	Wind turbine/1 kWh of electricity	Extraction raw materials, manufacturing turbine components and wind plant components, installation, maintenance and end-of-life disposal. Blades are 100% recyclable (shredding, incineration), foundation is 50% recyclable (crushing).	ILCD midpoint and ReCiPe endpoint	Climate change (onshore average) 6.5 g CO2-eq Climate change (offshore average) 9.35 g CO2-eq
Poujol et al., 2020 [84] and Boulay et al., 2017 [85]	Floating wind park/1 kWh of electricity	Extraction raw materials, manufacturing turbine components and wind plant components, installation, maintenance and end-of-life disposal. Blades are not recyclable.	ILCD, CED AWARE ReCiPe 2016 midpoint (H)	Climate Change 22.3 g CO2 eq Resource Depletion 14.6 mg Sb eq Water Use 6.7 L Marine Ecotoxicity 6. 91 g 1,4-DBC eq Air Quality 25.2 mg PM2.5 eq CED Renewable 22.4 kJ CED non-Renewable 429 kJ
Vestas, 2015 [86]	Wind park/1 kWh of electricity	Extraction raw materials, manufacturing turbine components and wind plant components, installation, maintenance and end-of-life disposal. Blades at EoL treatment: 50% incineration, 50% landfill	CML 2013, GaBi DfX	Abiotic resource depletion (elements) 0.06 mg Sb-e Abiotic resource depletion (fossils) 0.10 MJ Acidification potential 32 mg SO2-e Eutrophication potential 3.7 mg PO4-e Freshwater aquatic ecotoxicity potential 62 mg DCB-e Global warming potential 7.2 g CO2-e Human toxicity potential 1427 mg DCB-e Marine aquatic ecotoxicity potential 793 g DCB-e Photochemical oxidant potential 3.9 mg Ethene Terrestrial ecotoxicity potential 4.1 mg DCB-e
Al-Behadili et al., 2015 [87]	Wind turbine/1 kWh of electricity	Extraction raw materials, manufacturing turbine components and wind plant components, installation, maintenance and end-of-life disposal. Blades are not recyclable.	N/a	SO2 0.0271291 g NOX 0.0382147 g CO2 10.420169 g N2O 0.0001474 g CH4 0.0001065 g NMVOC 0.0003469 g CO 0.0112237 g
Chipindula et al., 2018 [88]	Wind turbine/1 kWh of electricity	Extraction raw materials, manufacturing turbine components and wind plant components, installation, maintenance and end-of-life disposal. Blades and foundation are not recyclable.	Impact 2002+ midpoint and CED, SimaPro	Onshore/Shallow-water offshore/Deep-water offshore: Carcinogens 29.64/89.93/79.54 kg C2H3Cl-eq Non-carcinogens 23.11/59.05/62.51 kg C2H3Cl-eq Respiratory inorganics 0.90/2.65/1.96 kg PM2.5-eq Ionizing radiation 4663/10,197/8838 Bq C14-eq Ozone layer depletion 4.64 × 10−5/9.59 × 10−5/9.76 × 10−5 kg CFC11-eq Respiratory organics 0.217/0.777/0.275 kg C2H4-eq Aquatic ecotoxicity 65,200/178,164/181,041 kg TEG Terrestrial ecotoxicity 2627/67,364/66,740 kg TEG Terrestrial acid/nutria 8.4/22.91/19.09 kg SO2-eq Land occupation 6.86/21.32/15.44 m2 Aquatic acidification 2.65/7.72/6.47 kg SO2eq Aquatic eutrophication 0.40/1.01/1.29 kg PO4 P-lim Global warming 440/1144/648 kg CO2eq Non-renewable energy 6578/16,115/10,930 MJ Mineral extraction 223/590/702 MJ
Alsaleh and Sattler, 2019 [57]	Wind turbine/1 kWh of electricity	Extraction raw materials, manufacturing turbine components and wind plant components, installation, maintenance and end-of-life disposal. Blades and foundation are not recyclable.	TRACI, SimaPro	Ozone depletion 2.69 × 10−9 kg CFC-11-eq Global warming 1.80 × 10−2 kg CO2-eq Smog 1.71 × 10−3 kg O3-eq Acidification 1.04 × 10−4 kg SO2 –eq Eutrophication 9.78 × 10−5 kg N eq Carcinogenic 1.78 × 10−9 CTUh Non-carcinogenic 5.88 × 10−9 CTUh Respiratory effects 1.46E−05 kg PM2.5eq Ecotoxicity 1.83 × 10−1 CTUe Fossil fuel depletion 2.02 × 10−2 MJ Cumulative energy demand 1.07 × 10−1 kWh Water depletion Index 9.24 × 10−7 m3
Guezuraga et al., 2012 [89]	Wind turbine/1 kWh of electricity	Extraction raw materials, manufacturing turbine components and wind plant components, installation, maintenance and end-of-life disposal. Blades are incinerated; the foundation is landfilled.	CED, GEMIS	2.0 MW-geared/1.8 MW-gearless CO2-eq 9.73/8.82 g Cumulative energy demand 0.654/0.645 kWh
Stavridou et al., 2019 [90]	Wind turbine/1 kWh of electricity	Extraction raw materials, manufacturing turbine components and wind plant components, installation, maintenance and end-of-life disposal. Blades are incinerated, the foundation is landfilled.	CED, GEMIS	Steel tower/lattice tower Cumulative energy demand 0.48/0.33 kWh
Lefeuvre et al., 2016 [91]	Blades, 1 kg of carbon fiber reinforced polymers (CFRPs)	The process of pyrolysis only	ReCIPe Midpoint (H), SimaPro	Global warming potential −0.08 kg CO2 eq
Li et al., 2016 [92]	Blades, 1 t of CFRPs	CFRP treatment routes: landfilling, incineration, recycling and landfilling, recycling, and incineration	EcoInvent data	Landfilling/incineration/recycling and landfilling/recycling and incineration Global warming potential 24/2011/-378/750 kg CO2-eq.

).

Garrett and Rønde [82] performed an LCA for a 50-MW wind park. The functional unit was 1 kWh of electricity produced. GaBi DfX software and primary data from Vestas were used. The analysis included all stages for the manufacturing and transportation of raw materials, turbine and wind plant components, as well as maintenance and end-of-life disposal. The study used data on primary fuel consumption collected by Vestas for truck and sea vessel transportation of turbine components. The transportation distance corresponded to the one that is part of Vestas’ supply chain. Turbine recyclability (in percent turbine mass) was estimated to be between 81% and 85% (mainly metals), depending on the class of the turbine. In modeling, an avoided impact approach was used.

Bonou [83] used the ILCD method for the assessment of the environmental impact of two types of wind power plants (onshore and offshore) from the extraction of raw materials to the EoL. The primary data were collected from four representative European power plants with state-of-art technology provided by Siemens Wind Power. The functional unit was 1 kWh. The EoL of the power plants consisted of the management of construction and demolition wastes. The recycling of turbine blades and foundation was included. The recycling process for composite blades was mechanical shredding and incineration in cement production. The recycling process for the cement foundation was crushing with the positive output of crushed gravel. The results in physical units were obtained only for the impact category “climate change”, and they indicated the preference of land-based wind parks.

Poujol [84] studied a floating 24-MW offshore wind farm’s life cycle from “cradle-to-grave”. The multicriteria approach was used for LCIA. It is based on the combined performance of ILCD, CED, and ReCiPe 2016 MidPoint method (H). An additional parameter of “water use” was estimated according to the AWARE method [85]. The functional unit was 1 kWh of electricity. The recycling technology was not specified. Presumably, this is the usual recycling of metals and landfill for composite blades. The contribution of the EoL stage to all the adverse effects was quite large due to the need to use diesel-powered marine vehicles for decommissioning a wind farm located 16 km from the coast.

In a report of Vestas [86], the environmental impacts associated with the production of electricity from a 50-MW onshore wind plant were studied using a cradle-to-grave LCA. The functional unit was 1 kWh of electricity. At the EoL stage, it was assumed that metals are recycled (85–87% of the total turbine mass), and composite blades are incinerated by 50% and landfilled by another 50%. Blades and foundation treatment did not bring avoided environmental impacts. An important distinguishing feature of this work is the high certainty of data on all processes. It included the process of transporting equipment to the installation site and the recycling site and the process of connecting to the grid. The estimated transportation of the turbine components ranged from 50 km for the base to 2200 km for the blades; the transportation distance to the recycling site was assumed to be 200 km.

Al-Behadili and El-Osta [87] calculated the emissions of a 1.65-MW wind turbine over its entire life cycle (including EoL). The functional unit was 1 kWh of electricity. An important feature was a fact that emissions related only to energy consumption were taken into account. The initial data were obtained from literature and taken in an averaged form. EoL treatment technologies were not specified, but from the context of the paper, one can conclude that it was recycling of the metal parts of the turbine and the landfilling of the blades. Transportation distance was not specified.

Chipindula [88] performed a classic LCA for several types of turbines: medium ground-level power (1, 2, and 2.5 MW), powerful offshore in shallow water (2 and 2.5 MW), and offshore in deep waters (2.5 and 5 MW). All intermediate and final calculated data were averaged over the class of turbines. The functional unit was 1 kWh of electricity. EoL treatment involved the recycling of metal components of the turbine in the range from 55% (for aluminum) to 90% (for iron, steel, and copper) and 100% disposal of composite parts and a concrete base. Transportation of turbine components to the installation site was taken into account, and the maximum distance was assumed to be 10,000 km. Transportation to the place of recycling was not taken into account.

The results of the study show that although offshore wind parks have a higher capacity factor (45–47% compared to 35% for onshore wind turbines), in most categories, they produce more significant adverse impacts. This is precisely due to the contribution of the EoL stage, in which concrete foundations are left in the ground.

Alsaleh and Sattler [57] performed a life cycle inventory and assessment of the various phases of the 2-MW Gamesa onshore wind turbine life cycle in terms of TRACI impact categories with SimaPro Software version 8.3.2. The study considered RES recycling (OLR) with 98%, 90%, and 50% recycling rates for metals, plastic, and electronic components, respectively. Fiberglass (blade material) and lubricants were considered 100% landfilled. The most notable result of the study is the conclusion regarding the environmental friendliness of transporting new turbines to the installation site: the impact of truck transport for the distance 656 km is, in most cases, comparable or even greater than that of transoceanic ship transport, which was 8325 km. The effect of transporting the used parts of the turbines to the place of recycling or landfill was not taken into account.

Guezuraga [89] used GEMIS software for modeling the entire life cycle of two types of wind turbines: a 2-MW turbine with a gearbox and a 1.8-MW turbine without a gearbox. The paper assumed that the 2-MW turbine was transported for a distance of 2700 km, and the 1.8-MW turbine was transported for a distance of 1100 km by truck. The transportation distance to the place of recycling or landfill was not indicated. Only energy-related impact categories were considered. Recycling of stainless steel, cast iron, and copper was considered, whereas epoxy, plastic, and fiberglass were incinerated. The concrete foundation was 100% landfilled. Comparing the results for the two turbines, the authors concluded that the impact of the 1.8-MW turbine without gearbox was a little less.

The authors of [90] compared energy demand for the entire life cycle of 2-MW onshore wind turbines with a tall (76.16 m) tower. The study compared the ecological impact of traditional steel and innovative lattice towers. A lattice tower needs around 35% less steel, and it has an almost 33% lighter foundation, which gives a significant advantage for transportation and construction. GEMIS software was used for modeling total energy demand. The transportation for 240 km by truck and 1020 km by ship to the location of installation was considered. EoL treatment included recycling of metals with 5–10% loss, the 100% incineration of epoxy, fiberglass, and plastic, and 100% landfill of the concrete foundation. The distance to the place of processing, incineration, or landfill was not indicated. The study concluded that the turbine with a lattice tower was 32% less impactful on the environment in terms of CO₂ emissions.

Of the studies modeling LCAs of composite materials, only a few can be distinguished. For example, in [91], recycling by pyrolysis of carbon fiber reinforced polymers (CFRPs) was investigated with an LCA according to the ReCIPe Midpoint (H) 1.10 method. The functional unit was 1 kg of CFRP waste. It was estimated that recycling can avoid −0.08 kg CO₂ eq. due to the recovery of methanol and ethyl acetate.

In [92], a limited version of LCA was carried out for several possible options for handling CFRP waste. Only CO₂ emissions were taken into account. The models in this study were based on hypothetical CFRP treatment routes because exact facility locations were not identified. “Gate-to-grave” models have been developed for CFRP waste treatment by landfilling and incineration, beginning at the point of waste collection and including waste processing (disassembly, shredding), transport, and waste treatment (landfill, incineration). For recycling, a “gate-to-gate” approach is taken, which includes the production of composite materials from recycled carbon fibers (r-CF) and the use and/or disposal of other recyclate materials. Maintenance and facility construction is also not included in the LCI boundaries.

The results demonstrated that landfilling produced minor GHG emissions (24 kg CO₂eq./t CFRP) due to the inert nature of CFRP waste; incineration resulted in the greatest net GHG emissions (2011 kg CO₂eq./t CFRP) from the combustion process. Energy outputs from incineration are assumed to displace the electricity and natural gas-fired heat generation, which gave a credit (−1041 kg CO₂eq./t CFRP). Mechanical recycling with landfilling of the coarse recyclate fraction and displacement of GF production resulted in a net global warming potential reduction of 378 kg CO₂eq./t CFRP.

The results of a comparative analysis of environmental assessments of the life cycle of wind turbines are summarized in

Table 6. Environmental impacts of different life cycle stages of wind plants (for 1 kWh).

Impact	P+O+R 4	Blades Recycling Potential	P+O 5
	[57,82,83,84,86,87,88,89,90]	[91,92]	EcoInvent
Climate Change, g CO2 eq	6.5 [84]–10.42 [87]	−0.01088 [91]–0.102 [92]	95.7
Resource Depletion, mg Sb eq	0.06 [86]–14.6 [83]		11.692
Abiotic depletion (fossils), MJ	0.02 [57]–0.1 [82,86]		1.4744
Water Use, L	0.0009 [57]–6.7 [83]		40.842
Marine Ecotoxicity, g 1,4-DBC eq	6.91 [83]		179.95
Air Quality, mg PM2.5 eq	14.6 [57]–25.2 [83]		203.25
CED, kWh	0.107 [57]–0.654 [89]		1.476
Acidification potential, mg SO2-e	27.1 [87]–104 [57]		540
Eutrophication potential, mg PO4-e	3.7 [86]		231.48
The photochemical oxidant, mg Ethene	3.9 [86]		39.714
N2O, g	0.0001474 [87]		0.003353
CH4, g	0.0001065 [87]		0.26812
NMVOC, g	0.0003469 [87]		0.47615
CO, g	0.0112237 [87]		0.6853
Carcinogens, CTUh	1.78 × 10−9 [57]		1.5899 × 10−8
Non-carcinogens, CTUh	5.88 ×10−9 [57]		1.0176 × 10−7
Land occupation, m2	6.86/21.32/15.44 [88]		0.012118
Ozone depletion, kg CFC-11-eq	2.69 × 10−9 [57]		8.1088 × 10−9

⁴—Minimum and maximum values are given in single units of measurement from sources [57,82,83,84,86,87,88,89] of Table 5. Data from [88] were not taken into account since they are 3–4 orders of magnitude higher than data from other sources, which most likely indicates significant differences in the choice of primary data. The exception was the data on the land occupation category since they were not found in other sources, and it was impossible to compare them with any other results. ⁵—Data from EcoInvent for Vestas 2 MW wind turbine (global).

. The following conclusion can be drawn: increasing the capacity of wind turbines by increasing its height and the span of the blades from an environmental point of view is unreasonable since transporting bulky components of the turbine to the installation and recycling site or landfill contributes too much to the overall ecology footprint. For the same reason, it is inappropriate to build offshore wind farms far in deep waters. Contrary to the findings of [93], we did not find solid evidence that the environmental impacts of onshore wind farms are higher. This situation can only change if environmentally friendly modes of transport (for example, electric trains) are used. Overall, these findings are in accordance with findings reported by Wang et al. [93]. The technologies for recycling the blades need further development. So far, it can be stated with a degree of certainty that incinerating the blades of wind turbines is an environmentally unacceptable alternative to their disposal. The development of technologies to reduce the weight of the tower for a wind turbine, for example, by improving its design, has significant prospects from the environmental point of view.

A major limitation of our study is the uncertainty induced by parameters of technological, spatial, and temporal nature in LCA models of wind and solar plants. For example, recent research [94,95] suggests that there is great variability in results within sets of wind turbines with similar nominal power output. Nevertheless, we can still state several important common principles that can be applied to the eco-design on energy systems based on RES. This is particularly important when investigating new possibilities for the development of renewables in countries where the issues of eco-design have not been properly addressed yet.

5. Applications for Eco-Design of Energy Systems and Energy Policy: The Case of Renewable Energy in Russia

Therefore, despite the incomplete data and the presence of significant uncertainties, the available studies of the entire life cycle of solar and wind energy plants allow us to draw several general conclusions regarding the eco-design of their production systems. Firstly, recycling is much more preferable from an environmental point of view than disposal, so it must be foreseen (including in the organizational and regulatory aspects) at the earliest stages of development of the renewable energy technologies. Secondly, when choosing a location for solar and wind energy plants, it is necessary to take into account not only the climatic and infrastructural conditions but also the distance from the place of the proposed location of the recycling enterprises.

Let us examine how this conclusion can be applied to the improvement of the eco-design of solar and wind energy projects in Russia. The literature review shows that there has been little discussion on the problem of RES capacity recycling in this country [96]. A reasonable explanation of this knowledge gap is an initial stage of development of wind and solar energy. To date, the most developed type of renewable energy in Russia is solar. In 2019, the volume of electricity generated by solar stations in Russia amounted to 850.38 MWh, while wind stations contributed only 209.84 MWh. Other types of renewable sources are represented by small hydropower plants (48.9 MWh), biogas plants (28 MWh), biomass and waste power plants (43.5 MWh), and landfill gas (9.6 MWh).

The system of government incentives for renewable energy in Russia stimulates not only the construction of new renewable energy plants but also the development of national manufacturers of solar panels and wind turbines. Due to these incentives, Hevel (the largest producer of silicon heterostructured solar modules in Russia) and NovaWind JSC (the Rosatom division responsible for wind energy projects) are already successfully operating on the Russian market.

Hevel currently manages 1145.5 MW of solar projects that are scattered throughout southern Siberia, the Volga region, the North Caucasus, and the Far East (Figure 5). The production unit of the company is located in the city of Novocheboksarsk, the Republic of Chuvashia. Thus, the distance between the manufacturer and the most remote solar energy facility is more than 7500 km.

At the time it was launched in 2015, the plant’s production capacity was only 90 MW per year; by 2019, it had already increased to 300 MW per year. The company plans to implement several large solar projects not only in Russia but also in Kazakhstan and further increase production to strengthen its position in the market and reduce production costs due to the economies of scale, learning-by-doing, and learning-by-researching [97,98]. Following this development strategy will likely lead to the construction of an increasing number of solar power plants, including ones at a considerable distance from the place of direct production.

Such an approach in the future can significantly complicate the development of the recycling of solar modules, making it environmentally and economically impractical due to the need to transport old modules over long distances. A more feasible solution, which allows optimizing the design of the production system even at this stage in the development of solar energy, could be the development of a network of Hevel subsidiaries in the regions with the greatest natural and infrastructural potential for solar generation. Furthermore, it is necessary to begin the development of technologies for the production of thin-film modules, the environmental friendliness of which, over the entire life cycle, exceeds the ecology parameters of the silicon modules.

NovaWind JSC was founded in 2017 as a system integrator of all wind projects launched at this point in Russia. The technology partner of the company is Dutch company Lagerwey. The production units of the company are located in the city of Volgodonsk, Rostov Region. Shortly, the company plans to reach a production capacity of 96 turbines per year. The first Adygea Wind Farm project, with a total capacity of 150 MW (60 wind turbines), was successfully launched in early March 2020. In addition, the implementation of several large-scale wind projects in the Stavropol and Krasnodar territories in the Rostov and Volgograd regions is planned for the next two years. All these regions are part of the Southern Federal District and border each other, so the logistics of wind projects can be considered optimal, from both economic and environmental points of view.

The first NovaWind JSC wind projects include the construction of wind farms using L100 direct-drive wind turbines with a capacity of 2.5 MW. The program is already under development for the start of the production of turbines with a high power of 4.5 MW. Since the negative impact of powerful turbines on the environment throughout the life cycle is greater than the impact of medium-power turbines, when developing such projects, special attention should be paid to the transportation of the large-sized parts of the turbines, as well as to ensure the possibility of their subsequent disposal.

6. Conclusions

Even though assessing the environmental impact of the life cycle of the fastest-growing types of renewable energy is a popular topic in scientific literature, it has still not been studied enough. The particular difficulty is the study of the recycling of obsolete solar and wind power equipment. The first reason for this is the complexity of the methodological approach, which involves the distribution of all the positive effects created in the recycling process to earlier stages of the life cycle. The second reason is the great uncertainty of the current estimates, both due to the lack of primary data and because of the variety of logistic solutions for supply chains.

This study systematizes the most recent and relevant LCA results and identifies, based on the comparison, the patterns that work under any assumption about the technology and organization of the recycling process. The paper concludes by arguing that the contribution of further improvements in PV recycling technologies to environmental impacts throughout the entire life cycle is insignificant. Therefore, it is more beneficial to focus further efforts on economic parameters, in particular, on achieving the economic feasibility of recycling small volumes of PV waste. In the case of wind power, broadly translated, our findings indicate that the issue of transporting bulky components of wind turbines to and from the installation location is critical for improving the eco-design of the entire life cycle.

These patterns can now be the basis for the design of production systems for generating electricity and thereby contribute to the development of a circular economy. It is important to note that those countries that are still at the beginning of the path of renewable energy development (for example, Russia) can take into account the positive and negative experiences of the leading countries and optimize their production systems at the very early stages of their development.

The main practical conclusion that can be drawn from this study is that in order to improve the environmental performance of solar and wind power plants, Russia needs to develop accurate data centers for the forecasting of waste flow. It will help to elaborate a reasonable proactive strategy and optimize the number and locations of recycling plants.

Future research should further develop and confirm these initial findings by monitoring the progress in RES environmental performance on each stage of their life cycles. In addition, the comparison of less developed RES technologies with an LCA “cradle-to-grave” approach might prove an important area for future research.