Collecting EOL LIBs is the first step toward bridging LIB consumption and the reclamation process. Currently, collection is a major challenge, and only a small amount of LIB is able to be recovered for recycling. The U.S. Department of Energy has estimated that the current collection rate can be as low as 5% (Simmons, 2021), while also recommending a long-term objective to achieve a recycling rate of 90% across all LIBs by 2030 (Granholm, 2021). Efforts toward a high LIB recycling rate can be compared to efforts to recycle lead acid batteries, which has been a nationwide success in the U.S., achieving recycling rates of over 99% (National Recycling Rate Study, 2022). However, currently the U.S., there is no federal regulation that promotes the recycling of LIBs. Some states have implemented battery recycling laws. Most notably California has passed the Responsible Battery Recycling Act of 2022 (LegiScan, 2023) to create battery producer-funded stewardship programs for collecting and recycling batteries. In the EU, a directive has been established to promote the collection and recycling of LIBs in member states (European Parliament of the Council, 2006) with a goal to reach 45% LIB recycling by 2016. The results have been mixed, as not all member states have been able to achieve the targets of the directive. Neumann et al. and Winslow et al. provide a more detailed summary of these recent global policy initiatives (Winslow et al., 2018; Neumann et al., 2022).

The CE market has traditionally dominated LIB usage, which is the leading cause for low collection rates. Collecting CE LIB is a challenging process compared to the EV and BESS LIBs because, in addition to mixed legislative policy success for implementing collection programs, no strong infrastructure exists for systematic collection. In the U.S., companies such as Call2Recycle have developed nationwide collection programs that partner with various retailers to provide collection sites for batteries. Despite being arguably the largest battery collector in North America, Call2Recycle reported about 1,000 tons of LIBs collected in 2021 (Annual Report, 2019), which is just a small fraction of the estimated LIBs available (48,500 tons; Gaines et al., 2021) to be collected and recycled. Therefore, it is critical that a number of technological, policy, and infrastructural initiatives are developed to address the gap in collection.

With the rapid growth of both EV and BESS sectors, a shift in the relative contributions of these market segments to EOL LIBs is forthcoming. The total volume of collectible domestic EOL EV and BESS is expected to exceed the volume of collectible EOL CE batteries in 2027 and exceed 1MM tons by 2030 (LaMotte and Clare, 2022). These markets have an easier collection process. For one, the average battery size is larger, leading to the handling of fewer batteries for an equivalent weight processed. For EVs, collection can occur at centralized locations, such as dealerships (AXIOS, 2022) or repair and maintenance shops (Li, 2022), where the responsibility of proper battery disposal is transferred to the service provider or contractor, instead of to the consumer. For BESS, the majority of LIBs will be handled by larger corporations or licensed contractors (Salim et al., 2019). These would also provide an opportunity to establish better disposal and collection processes over the CE market sector.

Battery sorting is a critical component of the battery recovery process, allowing for better aggregation of similar batteries, more efficient transportation, and the development of more effective processes for battery recycling. Collectors, especially for CE LIBs, will send collected mixed battery waste streams to sorters for further processing. LIBs sorting can be separated into three steps: isolating LIBs from other batteries or waste (Bebat, 2022a; How to Identify Different Types of Battery Chemistries, 2022); diagnosing the LIBs based on state-of-health (SOH) factors, such as remaining useful life (RUL) and state-of-charge (SOC), to decide if the LIBs can be repurposed for secondary use or recycling (Pagliaro and Meneguzzo, 2019; Martins et al., 2021; The Equation, 2022); and sorting the battery based on different major properties, such a cathode chemistry (Fricke, 1999; Larouche et al., 2020; Call2Recycle, 2022). An example breakdown of these processes is shown in Figure 3. Currently, sorting by type is widely used as the primary form of sorting with more development being done on SOH and chemistry-based sorting for LIBs.

Considering the difference in the recovery processes among various types of batteries, sending the appropriate battery types to their corresponding recyclers is the first and arguably most important step. Sorting by battery type also benefits further storage and transportation since lithium-containing batteries having special transportation (Huo et al., 2017) and storage (CTS, 2022) requirements that tend to be more costly than other batteries. For LIB and non-LIB separation, manual sorting processes dominate the market. Trained personnel visually identify and physically sort batteries that are placed in front of them via a conveyor or sorting table (Eco Batt, 2021; Bebat, 2022a,b; Sorting Batteries, 2022). Such a process relies heavily on the training and capability of personnel (in addition to overall availability of labor) to identify and sort batteries accurately and efficiently. More recent developments in battery sorting automation rely on vision-based technologies and artificial intelligence to separate batteries based on their optical characteristics, such as label or shape (Eco Batt, 2021; Snapshot, 2022). This process is claimed to have higher sorting efficiency and higher accuracy. Although the vision technology helps with collecting data on batteries and feeding them into a machine learning algorithm, which further optimizes the efficiency and accuracy of sorting, the technology has some limitations. Refind Technologies has designed a line to identify all cylindrical batteries up to D size and sort all 9 V batteries into four classes: alkaline, NiMH, NiCd, and lithium (Eco Batt, 2021; Recycling Product News, 2022), while Opti-Sort can only sort cylindrical batteries (Harper et al., 2019). Also, a large amount of manual pre-sorting (Bebat, 2022a) can be required before entering the vision sorting machine to remove non-conforming batteries, which further reduces the efficiency of the overall process. Li Industries recently developed a Smart Battery Sorting System that leverages vision, chemical measurements, machine learning, and industrial internet of things (IIoT) technology to automatically sort a much wider range of battery types, shapes, and chemistries (Call2Recycle, 2022; Submission, 2022). However, vision technologies alone are limited in their ability to identify batteries, especially when the batteries are not packaged uniformly or are damaged. These issues could be alleviated by more standardized disposal and collection policies (Call2Recycle, 2017; Team, 2019). Overall, continued development is needed to fully deploy an automated solution into the sorting industry to reduce pain points related to labor cost, accuracy, and efficiency.

Not all discarded LIBs are truly at end of life. Although they may no longer be optimal for continued use in their original application, many LIBs qualify for second life use based on their SOH, since different applications have varied criteria for performance. For example, the LIBs retired from the EV can be reused for BESS (Lyu et al., 2019; Battery, 2021). However, due to the inconsistency of degradation among cells in the pack, the SOH of each cell may be different. Thus, the diagnosis or prediction of SOH and remaining useful life (RUL) can be applied to determine if LIBs need to be recycled or whether they can be reused. The prediction approaches, according to the mechanisms they used, can be classified as model-based methods and data-driven methods (Berecibar et al., 2016; Li et al., 2019). Model-based SOH and RUL simulate the behavior (e.g., electrochemical behaviors and thermal behaviors) of batteries with numerical models, and thus predict the SOH and RUL. Electrochemical models that solve the mass and charge kinetics with partial differential equations are typically employed (Ramadesigan et al., 2012). Electrical models are also widely used with electrical-circuit analogs (Nejad et al., 2016). Data-driven methods for SOH and RUL prediction require large sets of degradation-related data rather than an accurate degradation model, thus relying heavily on the experimental data that captures the historical observations of the batteries (Hu et al., 2015; Xiong et al., 2018).

For both model-based and data-driven predictions, the most commonly used diagnostic method is physically making an electrical connection to the LIB and applying various electrochemical assessments. This group of diagnostic methods can be separated into the static and dynamic methods, while the static method is a one-time measurement and not related to the working status, dynamic diagnosis focuses on the change of the parameters during the charging and discharging process (Xiaoyu et al., 2014). Table 2 shows different examples of parameters and methods used in determining SOH of LIBs and their advantages and disadvantages. Static diagnosis takes measurements when batteries are in open circuit. The diagnosis can finish within seconds, but the prediction error can be as high as 10% (Berecibar et al., 2016). Dynamic diagnosis measures the parameters while batteries are in operation. Such diagnosis needs to take measurements over a portion of or a whole operation cycle (which can take up to hours), but the prediction error tends to be lower (< 5%) within well-modeled working conditions (Waag et al., 2014; Berecibar et al., 2016; Lipu et al., 2018; Sarmah et al., 2019). The drawback of the physical contact measurement is that single-parameter sorting methods are not accurate while multiple-parameter sorting methods increase the complexity and time (Chen W. et al., 2019). Non-contact diagnosis using X-ray (Chen and Shen, 2017) or ultrasound technology has also been proposed and developed. Titan Advanced Energy Solutions has developed a new sorting technology that uses ultrasound to detect the LIBs SOH, which is claimed to be fast and accurate (Titanaes, 2022). Furthermore, analysis methods utilizing fuzzy logic, machine learning, or other computational techniques have been used to increase the accuracy of assessments (Xiong et al., 2018; Lyu et al., 2019).

A third step of sorting that can prove useful to downstream processing is cathode material sorting, which separates the LIBs into different categories by their cathode chemistry. Cathode sorting helps hydrometallurgical and pyrometallurgical recycling reduce the complexity of the process and increase the purity level of recycled materials (Larouche et al., 2018). The value of generated material from these processes is improved if, for example, sorting provides LIB waste streams with high Co or Ni content (e.g., LCO, NCM, NCA), when compare to LIB waste streams with higher Fe or Mn content (e.g., LFP, LMO), due to the relative value of these different metals (Larouche et al., 2020; Lander et al., 2021). Sorting is also necessary for more sustainable and cost-effective recycling technologies such as direct recycling, which requires relatively pure single cathode waste streams. Cathode materials can be retrieved from battery information databases if those chemistries are known for particular battery models, which can be more common in the EV and BESS LIB market due to the lower number of battery models available. For CE LIBs, the identification of different models and correlation to a database of chemistries may prove to be more challenging. Overall, visual identification methods could be improved with the better standardization of labeling or if other identification technologies such as RFID and QR code or barcode tags are used (Preradovic and Karmakar, 2010; Gaines, 2018; Zheng et al., 2021). Other available technologies for identifying cathode chemistries include both invasive and non-invasive spectroscopic techniques (OnTo-Technologies, 2014; Submission, 2022), which have been recently developed in automated sorting systems.

Pretreatment is essential to all types of the recycling process. The pretreatment includes disassembly, discharging, and shredding, comminution, or exposing of the LIB core materials. Pretreatment helps to reduce the safety issues by removing any residual energy stored in the LIB and improve consistency among starting materials that are entering downstream processes (Zhang et al., 2013). Different types of LIBs may require different levels of pretreatment to render the battery sufficiently safe or prepared for further processing.

For all types of recycling methods, the disassembly process is essential for certain types of LIB formats. The majority of EV and BESS LIBs, and some CE LIBs, are assembled into a pack containing multiple single LIB modules and/or cells. For ease of recycling, the packs can first be disassembled into modules, and modules are disassembled into single LIBs (Harper et al., 2019; Wu et al., 2022). The LIB packs can be large and heavy and include extra accessories (Arora et al., 2016; Xia et al., 2018; ZDNET, 2022), such as electronics, cooling systems, and various types of packaging (Harper et al., 2019). The large size and varying formats make packs hard to be directly processed by semi-automated or fully automated machinery. The constituent materials of the extra accessories can also introduce impurities during the recycling process, so separation and removal of these extra accessories is ideal (Zhang et al., 2013).

Disassembly can be manual (Blankemeyer et al., 2021; Martins et al., 2021), semi-automatic (Fan et al., 2022), or automatic (Zorn et al., 2022). Manual disassembly is common but not as efficient or safe (Waldmann et al., 2016). However, when there are a variety of pack and battery form factors and other design features, manual disassembly may be required. The semi-automatic disassembly process comprises both manual steps and some automated machine-assisted steps (Herrmann et al., 2014). Machines replace the worker to work on tedious and redundant tasks such as removing screws and moving modules out of the packs (Kay et al., 2022). Semi-automatic processes are popular in industry, especially for larger and more standardized battery systems (Fan et al., 2022). Fully automatic disassembly, on the other hand, is an ideal standard as it reduces labor costs and improves efficiency. However, this mode of disassembly is challenging because of the lack of conformity among different LIB pack designs (Zorn et al., 2022). An automatic disassembly line can be expensive and may need to be adapted for each different pack design, thus reducing practicability for most applications except for those that have a low pack design variance (Zorn et al., 2022). Harper et al. (2019) provides further review on the recycling of LIBs with a focus on the EV sector, with discussion on the different design considerations and types of disassembly (Harper et al., 2019). The authors highlight the variation among battery shapes and the lack of standardization for battery cells, packs, or systems, emphasizing the challenges in all types of efficient disassembly, especially in attempts to automate disassembly processes.

The LIBs before or after disassembly need to be discharged before they can be safely processed further (Gratz et al., 2014; Chen W. et al., 2019; Sommerville et al., 2020). If undischarged LIBs are shorted during the recycling process, the LIBs can cause serious damage due to thermal events. This issue has been of particular concern in proper LIB waste management with LIB-related fires in waste management facilities becoming more common in the last decade (O'Connor, 2021). Discharging, or an equivalent process, seeks to reduce or eliminate this use by removing most of the residual, stored energy in LIBs. Common discharging methods includes resistive discharging through a solid electrical connection (Dewan et al., 2010; Huria et al., 2021), thermal discharging (Feng et al., 2018), and salt solution discharging (Lv et al., 2018). The LIBs can be discharged by connecting the anode and cathode with high-power resistors (Sommerville et al., 2020). Such discharging systems have become commercially available and are designed to accommodate various types of LIBs (such as different EV LIBs) (Dewan et al., 2010). This method can be limited by varying LIB designs and so can be hard to accommodate all incoming LIBs. Thermal discharging heats the battery at 100–500°C for a specific time (Sommerville et al., 2020). The degradation of LIBs components makes the LIBs self-discharging (Mo Seong et al., 2018). The method is compatible with pyrometallurgical recycling because it can be combined with the steps in the pyrometallurgical. In fact, in many cases, pyrometallurgical recycling may not require a separate discharge process at all (Wu et al., 2022). Salt solution discharging involves an ionic process that electrolyzes water to discharge LIBs (Nembhard, 2020). This method has been widely used for all types of LIB recycling. Some other processes seek to bypass a separate discharge process entirely. Shredding the LIBs under liquid nitrogen or inert atmosphere prevents rapid oxidation of the LIB when the electrode are exposed to air (Fan et al., 2020). Shredding with the application of a water spray can discharge the battery while shredding the battery. The water also provides a method of controlling the thermal discharge from the battery during this process, since the overall process is less controlled than other discharge methods but is preferred in many recycling operations due to lower processing costs (Harper et al., 2019).

Physical separation is essential to hydrometallurgical recycling and direct recycling and can be used in pyrometallurgical recycling as well. The separation processes rely primarily to mechanical methods to isolate various materials such as the current collector (Cu, Al), packaging materials (Fe, plastics), separator, and electrolyte from cathode and anode materials. The removal of these materials can improve the effectiveness of downstream processes, as well as reduce the amount of low value materials present in the processing stream. The powder generated from the physical separating line is called “black mass”, which contains the cathode and anode active material, binder [e.g., polyvinylidene fluoride (PVDF)], carbon, and some Al, Fe, Cu, and organic impurities (Punt et al., 2022).

Depending on the cell format of the LIBs, which can vary depending on application, the physical separation process is also different. Figure 5 shows several examples of the black mass generation process for common cell formats, such as polymer case pouch cells, Al case prismatic pouch cells, and steel case cylindrical cells (Full et al., 2020). The first step is crushing the discharged individual cells or small packs with high-power crushers, which downsizes the LIBs into smaller flakes (Shin et al., 2020). Prismatic cells which have heavier casing materials (Full et al., 2020) can first be separated by weight using a process like wind sifting. This sifting process to remove larger, heavier materials (such as larger Al casing) reduces issues downstream with shredding. A shredding process can mill the small flakes into a coarse powder (Ku et al., 2016; Wang X. et al., 2016; Ebin et al., 2018). For cylindrical LIBs with steel casing, the steel casing can be left in the coarse powder or removed by magnetic separators (Harper et al., 2019; Sommerville et al., 2020). The separators in LIBs can be removed by a second wind sifter (Vaško, 2015; Zheng et al., 2017), which applies relatively weaker air flow when compared to the first wind sifting. Then, the coarse powders are processed by sieving to remove larger-sized particles. The blocked materials can be further shredded into smaller particle sizes to obtain finer powder (Ku et al., 2016; Wang X. et al., 2016; Ebin et al., 2018). After the second shaking sieve, the majority of the black mass is separated from the rest of the materials, including Al and Cu current collector, as well as packaging materials. The black mass can be treated by furnaces around 300–500°C to remove the binder and electrolyte. The mixtures of Al, Cu, and packaging materials can be further isolated from each other from the black mass by a gravity separation (Vaško, 2015; Sommerville et al., 2020). The lighter Al and packaging materials are brought to one side by the application of airflow, thus separating from heavier Cu powders. The steel, Al case, separator, Al powder, Al polymer case, Cu powder, and black mass are collected and packaged individually. The black mass generated from physical separation is widely recycled by the hydrometallurgical or pyrometallurgical methods. The black mass with low impurity can also be recycled by direct recycling with the addition of cathode/anode separation steps, such as froth flotation (Kim et al., 2021; Verdugo et al., 2022).

After separation, hydrometallurgical methods use one or more aqueous solutions to leach the metals in LIBs, which is then followed by different stages of purification and precipitation to recover the transition metals and lithium. A variety of aqueous solutions have been tested and applied to leach the LIBs. Alkaline leaching includes ammonia-based systems, sodium hydroxide, while acid leaching typically utilizes inorganic strong acids, such as HCl, H₂SO₄, and HNO₃, or organic acids, such as citric acid, ascorbic acid, oxalic acid, and formic acid (Li et al., 2012; Ku et al., 2016; Gao et al., 2017; Huang et al., 2018; Du et al., 2022). After leaching, purification procedures are used to remove impurity ions. The purification methods include solvent extraction, chemical precipitation, and other separation methods like electrodeposition, electrolysis, adsorption, flotation, and ion exchange (Larouche et al., 2020). The solvent extraction method uses the difference in solubility of metal or metal compounds in aqueous or organic solutions. The chemical precipitation methods rely on the differences in the dissolution of metal compounds under various pH conditions (Du et al., 2022). After purification, the metals are precipitated using methods such as co-precipitation and sol-gel. Both methods adjust the proportion of metal ions in the leaching solution and use some precipitate agent to form elemental compounds. The co-precipitation method is limited to the inorganic acid leaching system while the sol-gel method can be applied to organic acid (Wu et al., 2022). Figure 6 shows several hydrometallurgical methods for different cathode materials. LCO generates materials such as Co(OH)₂ and Li₂CO₃ (Wang et al., 2011); NCM decomposes to Ni_1/3Co_1/3Mn_1/33(OH)₂ (or alternatively isolated Ni, Co, and Mn compounds) and Li₂CO₃ (Sidiq et al., 2022; Singapore Battery Consortium, 2022), and LFP can be broken down into FePO₄ and Li₃PO₄ or Li₂CO₃ (Li H. et al., 2017; Larouche et al., 2020). Those products generated by hydrometallurgical processes are Co, Ni, Mn, Fe compounds, and Li salt, which can be precursor materials but are typically not able to be directly used in the production of new LIBs. To make LIB precursor, sometimes additional processes are required. For example, the common products NiSO₄, CoSO₄ and MnSO₄ from hydrometallurgical methods need an additional precipitation or crystallization procedure to be ready for use in NCM synthesis (Ma et al., 2020).

Although pyrometallurgical recycling can also benefit from physical separation processes (e.g., by increasing the amount of higher value material entering downstream processes; Ruismäki et al., 2020) LIBs with less pretreatment and separation are also able to be directly processed. High temperature treatments smelt the LIBs to form a metal alloy. The organic components, such as binder, electrolytes, and separator, are burned off during the smelting process (Wu et al., 2022). Pyrometallurgical recycling can be categorized into three groups: direct roasting (Li et al., 2014), In situ reduction roasting (Zhou et al., 2021), and salt roasting (Makuza et al., 2021; Du et al., 2022; Wu et al., 2022). The direct roast reduces the metal oxide at >1,000°C with reducing agents. In situ roasting involves pyrolysis under a vacuum or inert atmosphere without additives, which uses a relatively lower temperature than direct roasting, between 800 and 1,000°C (Li et al., 2016; Xiao et al., 2017; Zhou et al., 2021). Salt roasting can be operated at even lower temperature <600°C, and reports have shown higher materials recovery rates (Wang D. et al., 2016; Dang et al., 2018; Fan et al., 2019). Because the products generated from pyrometallurgy are alloys and slag, further processing is needed to improve the value of the generated materials. Hydrometallurgical processing, using leaching and purification steps to isolate elemental compounds, is typically employed downstream of pyrometallurgical recycling (Mayyas et al., 2019; Dang et al., 2020).

Despite high energy use and higher emission rates, the pyrometallurgical process is popular in industry because of its simple procedures, high productivity, and its ability to leverage off existing mining procedures and equipment (Liu W. et al., 2019; Makuza et al., 2021). However, recently, the large energy cost and toxic emissions are making this process less attractive. Furthermore, pyrometallurgical recycling can efficiently extract those heavy elements like Co, Ni, Fe, and Cu, but the anode and organic components are not able to be recovered. To recover Li, extra steps need to be added which can result in higher costs and longer processing times (Dunn J. B. et al., 2022).

Unlike pyrometallurgical and hydrometallurgical recycling methods, which destroy the structure of cathode materials, direct recycling technology collects and purifies electrode material without destroying the structure (Figure 7). Various forms of physical separation can be used with direct recycling, although due to higher purity requirements for separation, some processes specific to direct recycling have been developed to better isolate the electrode materials from the rest of the LIB. After disassembly, deactivation, and some other processing to open cells, such as shredding, direct recycling methods focus on separating the electrode materials from the current collector by removing the binder (Ross, 2020; Zhan et al., 2020). The most common binder is the nonaqueous organic material PVDF. Aqueous binders, such as styrene-butadiene rubber (SBR) and carboxymethyl cellulose (CMC), gained popularity recently as well (Courtel et al., 2011; Zhang et al., 2021). Most binders can be removed by heat treatment or solvent dissolution. For example, PVDF decomposes to C/CO₂ and HF at temperatures of 400–550°C (Zhan et al., 2020; Fujita et al., 2021; Wang et al., 2021)and can also be dissolved by organic solvents, such as N-methyl-2-pyrrolindone (NMP), dimethylformamide, or the greener solvent Cyrene (Marshall et al., 2021). Aqueous binders can similarly be decomposed at elevated temperatures (Zhang J. et al., 2018) or washed in an aqueous solution (Courtel et al., 2011; Li J. et al., 2020).

After binder removal, electrode powder can be separated from the current collector by mechanical means (e.g., water sonication, mechanical washing, sieving, and air separation; Li X. et al., 2017; Sloop and Sloop, 2023). Those electrode powders contain cathode active materials, anode materials (normally graphite), and small amount of Al and Cu impurities from current collectors generated during the separation process. Studies have reported removal of these impurities using various methods, such as low-concentration alkaline solutions (Ku et al., 2016; Chen M. et al., 2019) or weak acid solutions (Fink et al., 2022) or mechanical, particle size, or density-based separation (Tsang et al., 2017). After purification, the anode and cathode material need to be separated; flotation and density separation are common techniques for this purpose (Shin et al., 2020; Folayan et al., 2021; Zhan et al., 2021). The flotation separates the cathode and anode based on hydrophobicity differences, since the carbon-based anode material is hydrophobic, and most cathode material is hydrophilic (Folayan et al., 2021). Density separation relies on separating the lower density graphite anode from the cathode materials (Wolf et al., 2021).

Upon completion of purification, high purity EOL cathode and anode materials are generated. However, these recovered cathode materials are typically lithium deficient and may have structural damage (Chikkannanavar et al., 2014; Yang et al., 2020). Likewise, the anode may have some residual lithium or other chemical or morphological changes (Sabisch et al., 2018). Regeneration processes are used to restore the performance of these electrode materials so that they can be redeployed in new LIBs. These regeneration processes are discussed in more detail below.

Some additional steps are required to bridge the output of the recycling processes with usable inputs in manufacturing processes. The products generated from pyrometallurgy are metal alloys and slags, like Co-Ni-Cu-Fe alloy, and Mn-rich slag. The products can be used in fields such as construction (Rämä et al., 2018). However, for the production of new cathode material, further processing is required. As mentioned above, typically hydrometallurgical processes are added to isolate elemental compounds, thus aligning the outputs of pyrometallurgical and hydrometallurgical to a similar set of materials. Following some additional processing, these products generated from hydrometallurgy can be considered as precursors of new LIB cathode synthesis.

Effectively, at this point in the process for hydrometallurgical and pyrometallurgical recycling, the downstream steps are identical to new cathode synthesis. For example, the metal precursors for NCM are mixed (e.g., sol-gel or co-precipitation) and subject to some preprocessing, further mixed with a lithium source, and then typically undergo a two-step synthesis to generate new cathode material. For standard synthesis procedures, the first step is around 400–550°C for 4–6 h, and the second step is around 800–1,000°C for 12 h (Xia et al., 2009; Hua et al., 2014; Duan et al., 2021). Table 3 outlines some example re-synthesis methods. Some recent hydrometallurgical methods also have proposed direct co-precipitation of the various cathode elemental compounds after dissolution and purification to produce mixed precursor that is ready to be mixed with a Li source to undergo re-synthesis (Duan et al., 2021; Ascend Elements, 2022). These processes are able to bypass the additional steps of precipitating precursor and then undergoing an additional mixing process. Upon completion of the synthesis steps, the new cathode material is ready to be used in new LIBs.

After the separation and purification procedure in direct recycling, purified but lithium-deficient cathode materials are generated. To recover the performance of the cathode, a regeneration process is required to relithiate and reverse any structural damage from previous electrochemical cycling. Relithiation methods includes hydrothermal (Shi et al., 2018; Xu et al., 2021), which requires elevated pressure and temperature; solid phase sintering (Nie et al., 2015; Li J. et al., 2020) which requires high temperature; ionothermal methods (Wang M. et al., 2020); and electrochemical methods (Yang et al., 2020). Although both hydrometallurgical/pyrometallurgical recycling and direct recycling require heat treatment procedures to generate functional cathode, the complexity and energy inputs are different, with direct recycling requiring a shorter heat treatment step on average (Table 3). The simpler procedures, shorter time, and lower temperatures make direct recycling potentially more environmentally sustainable and economically efficient.

The direct recycling of LFP can have additional advantages not observed in other cathode materials. Commercial LFP materials all come with conductive coating to increase the electronic conductivity and rate performance (Doeff et al., 2006; Oh et al., 2010). Also, nano-sized LFP materials have become popular recently and demonstrate superior performance compared to micron-sized LFP (Ding et al., 2010; Borkiewicz et al., 2015). During the pyrometallurgy/hydrometallurgy, the original carbon coatings are destroyed, and the generated precursors do not retain the original particle size (Chagnes and Pospiech, 2013; Zhang G. et al., 2019; Larouche et al., 2020). To produce commercial LFP from these recycled precursor components, the entire synthesis process must be redone from scratch. Shown in the Figure 8, ball milling (Gao et al., 2007) or nano-bead milling (Islam et al., 2015) is the most commonly used process to downsize precursors, during which 0.3-0.4 mm milling balls are added to the precursor mixture. The rotor run under high rotation speed to stir the precursor mixture with milling balls for 1–3 h. After milling to the nanoscale, the mixture needs to be dried by spray dryers, which preserves the original chemical and morphological properties (Gao et al., 2007; Rigamonti et al., 2020). During the synthesis process, the carbon coatings are generated from the incomplete burning of organic compounds under N₂, H₂/Ar, or other inert atmosphere (Belharouak et al., 2005; Mi et al., 2005; Konarova and Taniguchi, 2010). On the other hand, direct recycling has a non-destructive treatment of the cathode during the separation and purification process, which retains the cathode's original structure, morphology, and coatings or other modifications.

The downside for direct recycling in this case is revealed when considering the evolution of morphology from older EOL LIBs. These older LIBs may not have the same cathode morphology and thus modification to the morphology during regeneration may be required (Xu et al., 2019). Likewise, for other types of cathode materials, such as NCM, changes in preference for composition, dopants/coatings, and other modifications may create a gap between the materials produced from recycling and materials desired for new batteries. Various methods of upcycling have been proposed as a way to bridge that gap to either change the composition or modify the recycled cathode material to meet performance or manufacturing processing demands (Schiavi et al., 2021; Lin et al., 2022; Parikh et al., 2022; Zhang et al., 2022).

Although graphite anode material is not considered as valuable compared to cathode material, recovered graphite from EOL LIBs can have multiple applications if properly recycled. There have been recent developments on anode recovery methods as a way of improving the overall sustainability and economics of recycling (Zhang J. et al., 2019). Figure 9 shows several recycling pathways for graphite. In some cases, methods have been proposed to recycle EOL LIBs with an initial separation of anode and cathode (Li et al., 2012; Yang and Xi, 2015), while in other cases the entire LIB is processed and separation occurs later (Punt et al., 2022). When anode sheets are able to be separated from the rest of the components, the graphite direct recycled using washing procedures to free the graphite from aqueous-based binders (Courtel et al., 2011; Li J. et al., 2020). The separated graphite can contain impurities, including Cu (Punt et al., 2022) and Li salts (e.g., Li₂CO₃, Li₂O, LiF, ROCO₂Li, CH₃OLi, and Li₂O) (An et al., 2016).

Pyrometallurgical methods are typically unable to recover the graphite because of the initial high temperature processing that burns off organic material. The graphite that is derived from hydrometallurgical methods can be of high purity if the other LIB components are dissolved and isolated (Ma et al., 2019). As described above, direct recycling of anode when the cathode and anode are not initially separated requires a separation/purification process such as flotation or density separation (Folayan et al., 2021; Zhan et al., 2021) to isolate the anode from the cathode. The resulting recovered graphite can have a small amount of impurities including Cu, Al, Li salt, and some cathode materials (Chen et al., 2022). To further purify the graphite, the lithium and other metal components can be extracted by acid leaching (Chen et al., 2022; Wang et al., 2022).

Purified recycled graphite powder can be reused in LIBs through a regeneration process (Sabisch et al., 2018). The graphite can also be repurposed for other applications using upcycling methods (Liu et al., 2020). For example, recovered graphite has been shown to be able to generate graphene (Zhang Y. et al., 2018; Yang et al., 2019) and carbon sorbent (Du et al., 2022).

The choice of recycling technology is highly dependent on the economics of recycling costs and outputs. These economics, in turn, are related directly to the choice of cathode chemistry in the manufacturing of the LIB. Although Ciez and Whitacre found that NCM cathode is about 20% of the cost of input materials into a NCM cylindrical cell (Ciez and Whitacre, 2019; Schneider et al., 2019) and 8% of the total manufacturing cost, the recovered cathode material or its elemental constituents can be most of the revenue generated from recycling (Mayyas et al., 2019; Lander et al., 2021). Therefore, the profitability, and viability of the recycling operation, is tied to the cathode that is being processed. Lander et al. performed an analysis on this profitability (Figure 10) for hydrometallurgical, pyrometallurgical, and direct recycling operations in the U.S. and found a wide range of results based on the cathode material of the input LIB. Pyrometallurgical operations tended to be unprofitable regardless of the input material, while hydrometallurgical operations could be profitable with cathode compositions containing more cobalt and nickel. Direct recycling was profitable in all cases (Lander et al., 2021). Yang et al. (2021) compare the different recycling methods cost among different countries and present the overall revenue generated per kg cell recycled (Yang et al., 2021). The authors found that in general pyrometallurgical operations were not profitable, while direct recycling operations were. Similar results are also shown by economic comparison work from Yu et al. (2021).

As LFP becomes a more dominant cathode in the market (Lunde, 2022), recyclability of these LFP LIBs are increasingly becoming a concern. The issue with LFP as a cathode material is that it requires the same recycling effort for a lower overall value, since the material lacks high value Co or Ni (Table 4). The majority of the material value of LFP is in the production process and less so in the raw materials, due to the relatively low economic value of the main constituents (Li, Fe, P) (Wang T. et al., 2020). Both pyrometallurgical and hydrometallurgical processes have intrinsic challenges with processing. Pyrometallurgy is not feasible due to the lack of usable product, while hydrometallurgy can extract primarily Li as a somewhat valuable product (Beaudet et al., 2020). Therefore, only direct recycling is a truly economically viable technology for recycling LFP LIBs, barring any subsidies or other policy-related assistance.