Commodifying a carcinogen: Critical raw materials from arsenic-laden groundwater

Abstract

Arsenic is a potent carcinogen whose presence in groundwater poses a persistent global public health threat. While groundwater treatment mitigates arsenic exposure, it generates concentrated arsenic-rich by-products that are universally considered economic and environmental burdens. However, arsenic is experiencing a renaissance. Both the United States and European Union have classified arsenic as a critical raw material (CRM) due to the role of metallic arsenic [As(0)] in materials needed for digital infrastructure and clean energy systems. Bringing these two material flows together, we show that arsenic removed during groundwater treatment can be transformed into pure amorphous As(0) nanoparticles via a two-stage process of alkali extraction and selective reduction. The creation of valuable CRMs from carcinogenic treatment by-products is a potentially disruptive technology for the water sector that can alter global As(0) supply chains. Revenue from upcycled As(0) can also help improve the challenging economics of water treatment in marginalized arsenic-affected regions.

Toxic groundwater treatment waste is converted to valuable arsenic metal, a critical material for clean energy and digital tech.

INTRODUCTION

The contamination of groundwater by arsenic (As), a potent carcinogen, continues to be an urgent global public health issue (1). An estimated ~200 million people worldwide are potentially exposed to dissolved As, often present as arsenite [As(III)] or arsenate [As(V)], in groundwater at concentrations exceeding the World Health Organization provisional drinking water guideline of 10 μg/liter (1). The human health toll from the As crisis in poor, rural areas of South Asia is particularly catastrophic, having been called “the largest mass poisoning in human history” (2). Chronic exposure to toxic As levels in drinking water leads to a broad range of health impacts, including low intelligence quotient in children, neurodevelopment disorders, various cancers, and death (3, 4). Groundwater treatment via As sorption to iron (Fe) (oxyhydr)oxide particles is a globally established public health safeguard against As toxicity in drinking water. Today, groundwater treatment is employed effectively but simply transfers As from water to a concentrated solid sludge (Fig. 1). The disposal of this waste has been a long-standing regulatory, economic, and environmental challenge.

Fig. 1. Conceptual diagram of As removal from groundwater and valorization of As-rich groundwater treatment sludge.

Groundwater treatment is practiced widely around the world (top), whereas the valorization of As-rich treatment sludge (bottom), which is the specific focus of this work, is a novel approach. Unsustainable disposal methods typically follow sludge storage, leading to environmental As emissions and the loss of valuable resources. The valorization of As-rich sludge replaces current disposal methods and yields metallic As(0), a valuable critical raw material (CRM). The digital images in the center of the circle show a typical As-rich groundwater treatment sludge storage basin (top; ~3 m in diameter) and suspensions of As-rich sludge (left) and valorized metallic As(0) (right).

The management of As-rich waste, which can annually exceed hundreds of tons per plant (5), is one of the highest costs incurred by treatment plants (6) and has been a historic barrier to widespread implementation of small-scale treatment systems in resource-scarce regions (7). In high-income countries, landfilling is the most common sludge disposal approach (8). However, landfilling is a costly and unsustainable disposal method that yields unacceptable environmental As emissions (9) and the loss of valuable resources (10), including the critical nutrient phosphorus (P), which is abundant in groundwater treatment waste due to its typically high concentration in As-contaminated groundwater (11). In low-income areas, including those affected disproportionally by the As crisis in South Asia, more hazardous disposal practices are common, including open disposal directly to surface waters and soils (fig. S1), which perpetuates As pollution (8, 12, 13). While waste management practices vary depending on local regulations and available resources, As-rich sludge is universally viewed as an unwanted burden.

However, As is experiencing a renaissance as societal values and economic drivers shift. In 2023, the European Union (EU) classified As, specifically metallic As(0), as a critical raw material (CRM; fig. S2) due partly to its use in materials needed to power the rise of digital technology, including generative AI, and for the transition away from fossil fuels (14). Similar CRM classifications have been made in the United States (15), thus providing new legislative incentives to recover and upcycle waste-derived As. Valorizing groundwater treatment sludge containing toxic As(III) or As(V) to create valuable metallic As(0) is particularly strategic because metallic As(0) is the only As compound having trade data tracked in the EU, indicating an established market. The demand for metallic As(0) stems from its role in the production of many industrially relevant and high-value materials, including batteries, alloys, and As-bearing semiconductors, such as gallium or indium arsenide and arsenene (16, 17). Presently, the world’s supplies of metallic As(0) largely derive from a single country (China), creating risks of global supply disruption (16, 18). Furthermore, traditional methods of creating metallic As(0) from mining and processing of raw As mineral ore are ineffective and notoriously damaging to the environment, with estimates of >90% of mined As emitted to soils during the process (19). As a result, recent calls in the scientific literature have expressed an urgent need for innovation in metallic As(0) production and supply (19).

The requirements to improve As-rich sludge management and create local sources of CRMs present a compelling opportunity to redefine carcinogenic As as a commodity. Converting toxic As waste into a useful resource offers a path to close the loop of water treatment through circular upcycling. To date, the resource recovery potential of As-rich groundwater treatment sludge has been ignored due to a combination of (i) the absence of practical methods to upcycle As and other CRMs, such as P, from the sludge and (ii) the entrenched misconception that As is merely a liability with no intrinsic value. While research on As recovery from mining wastewater is growing (20–22), the relevance of this previous work is limited to a few countries that produce As-laden mining wastewater, in contrast to the far more widespread presence of As-contaminated groundwater, which affects numerous countries across every inhabited continent (1, 23, 24). Therefore, previous research on As recovery from mining wastewater cannot fundamentally resolve global As(0) supply vulnerabilities arising from the localized and concentrated nature of current As(0) production. In addition, As recovery from mining wastewater is inapplicable to geogenic As due to the substantially different chemical compositions of the two As sources, with mining wastewater containing much higher dissolved As concentrations and lower Fe and P levels than geogenic As sources (25, 26). Advancing new methods to valorize geogenic As sources is therefore essential.

In this work, we aim to develop a new chemical method to form valuable metallic As(0) from As-rich groundwater treatment sludge collected from a variety of different treatment plants. The two-stage upcycling process consists of extracting As from the solid sludge using alkali solution (extraction), followed by selective reduction of extracted As (refinement) via a common sulfur-bearing reducing agent, thiourea dioxide, yielding pure As(0) nanoparticles and nontoxic aqueous reaction products. In addition, the reductive formation of As(0) particles unlocks the recovery of aqueous P, which would otherwise be difficult to separate from aqueous As due to their similar charge and shape. The valorization of As treatment waste has potentially far-reaching implications for groundwater treatment operations and global As(0) supply chains. At the same time, and perhaps most importantly, As upcycling can help improve the challenging economics of water treatment in marginalized regions affected most profoundly by As contamination (1, 27, 28). This work sets the foundation for rethinking the role of As treatment as a dual service that provides safe drinking water and generates local sources of CRMs to meet legislative CRM production goals, such as those described in the EU’s 2024 Critical Raw Materials Act (29).

RESULTS

Separation of As and P via alkali extraction

The sludge produced from the removal of As from groundwater is a heterogeneous solid mixture of elements that consists primarily of Fe(III) (oxyhydr)oxides (>20 to 30% Fe) with various surface-bound species (table S1), including the most oxidized form of As, As(V) (up to >2.0 g As/kg; present as AsO₄ oxyanions), and P (up to >20 g P/kg; present as PO₄ oxyanions). A critical initial step for sludge valorization is to separate surface-bound As from the residual solids before reductive As(0) formation. We performed this extraction step by introducing the sludge to an alkali solution to desorb negatively charged species, yielding an As- and P-rich liquid phase that is easily separated from the residual solid sludge.

Alkali extraction with 1 M NaOH was highly effective at mobilizing As from treatment sludge (Fig. 2) due to a combination of processes induced by high OH⁻ concentrations, including increased charge repulsion between the negatively charged sorbent surface and deprotonated As(V) ions (AsO₄³⁻) and a reduction in As sorption sites arising from crystallization of the Fe(III) (oxyhydr)oxide sorbent (30). The addition of 1 M NaOH resulted in As extraction efficiencies exceeding 99% and aqueous As levels >250 mg/liter, depending on the sludge/liquid (S/L) ratio and sludge composition. For example, at the lowest S/L ratio (13 g/liter), 98 to 99% of solid-phase As was released from almost all sludges, resulting in minimal As bound to the residual Fe(III) solids. At an S/L ratio of 200 g/liter, the highest As concentration was measured in the extraction solution for all sludges (e.g., >250 mg/liter for 2.1_AerFe^IIOx), but some As (<30%) also remained bound to the residual Fe(III) precipitates. These results reveal a trade-off for sludge valorization. Lower S/L ratios can be applied if reuse of the residual Fe(III) (oxyhydr)oxides requires minimal As content (i.e., <20 mg/kg), which comes at the expense of lower extracted As concentrations, whereas higher S/L ratios are beneficial for creating the highest concentrations of extracted species but yield more residual solid-phase As.

Fig. 2. Extraction of aqueous As and P from groundwater treatment sludges.

Data are plotted for experiments using 1 M NaOH at solid S/L ratios of 13 to (A), 80 (B), and 200 g/liter (C and D). The symbols correspond to different sludge samples and the symbol shading scales with the initial As sludge mass fraction. The As or P extraction efficiency, defined as the extracted As or P concentration normalized by the total As or P concentration added as raw sludge, is given in percentage adjacent to select data points. Note the different y-axis scales for all panels. In this and subsequent figures, sample names indicate the initial sludge As content in gram per kilogram dry weight, followed by the As removal method used by the treatment plant (i.e., 1.2_FeCl₃ represents a sludge containing 1.2 g As/kg that was generated by FeCl₃ addition; further descriptions given in Materials and Methods).

An important benefit of alkali extraction is the corelease of P, which is a valuable CRM needed for global fertilizer production (11). Because of the higher P content of groundwater treatment sludge compared to As, alkali extraction resulted in aqueous P concentrations >1000 mg/liter for most sludges (Fig. 2 and fig. S3), although the extraction efficiency for P was typically lower than As (e.g., 81% As and 30% P extracted for the 0.9_AerFe^IIOx sludge at S/L = 200 g/liter). While As and P were both released during alkali extraction, the timescales of optimal P and As release differed substantially. In contrast to the behavior of As, the aqueous P concentration for most sludge samples peaked within the first 1 to 24 hours of extraction. For example, >60% of solid-phase P was solubilized from the 1.2_FeCl₃ sludge after 6 hours, which decreased to only 24% after 7 days. The decrease in aqueous P concentration with increased extraction time can be explained by the formation of an insoluble secondary Ca-P precipitate following P release, which was not observed for As likely due to the lower solubility of Ca-P phases compared to Ca-As phases (31). The reprecipitation of a P-bearing solid at later stages of extraction is consistent with the high pH and high aqueous P concentrations of the extraction solutions and the typically high fraction of Ca in the initial sludges (table S1), which creates favorable conditions to form Ca-P precipitates. Although aqueous P levels decreased with extended extraction times, both As and P remained at high concentrations after extraction, indicating a strong opportunity for corecovery if aqueous As and P can be separated subsequently.

Reductive As precipitation

Refinement of the alkali extraction solution (composition summarized in table S2) was performed by reductive As precipitation using thiourea dioxide (TDO), a low-cost and low-toxicity reductant commonly applied in the textile industry as a bleaching agent (32, 33). Using TDO is advantageous because of its enhanced reactivity under basic conditions, consistent with the extraction solution. Elevated pH destabilizes the TDO molecule, cleaving the C─S bond to yield nontoxic urea and redox-active sulfoxylate (fig. S4), which initiates As reduction (22, 34). While other chemicals can reductively precipitate As(0) (20, 35, 36), these reductants require acidic conditions, negating their applicability here.

The addition of TDO resulted in >99% removal of aqueous As from the extraction solutions in optimum conditions, with the efficiency increasing with the TDO/As mol ratio and aqueous As concentration (Fig. 3). For example, aqueous As in the extraction solution of the 2.1_AerFe^IIOx sludge decreased 99.3% from 300 to <2 mg/liter at a TDO/As mol ratio of 200. Simultaneously, the aqueous P concentration in this extraction solution remained constant at nearly 1700 mg/liter for all dosages of TDO. Similar trends in reductive precipitation of aqueous As with unaffected aqueous P levels were observed for the other sludge extraction solutions (Fig. 3 and fig. S5). The lowest aqueous As conversion efficiency was observed for the 0.2_AerFe^IIOx sludge (77% at TDO/As = 600 mol/mol) due to this sample having the lowest extracted aqueous As concentration, which decreases the efficiency of reaction with TDO (22, 34). Sludge upcycling is thus possible for solids with As contents ≤0.2 g/kg, but elevated As sludge levels >0.6 g/kg are optimal. This threshold value provides a design criterion for treatment plants implementing As recovery, regardless of the As removal method used by the treatment plant. The increased efficiency with increasing As sludge content also highlights the As upcycling potential of other types of intensely contaminated materials, such as mine tailings.

Fig. 3. Removal of aqueous As (colored bars) from extraction solutions via reductive As precipitation.

Data are plotted for experiments with different TDO/As mol ratios. The aqueous P concentration is given in open symbols. The residual aqueous As fractions are listed in % for selected experiments. The sludge sample names are listed above each bar chart and the bar chart color scales with initial sludge As mass fraction.

The different behavior of aqueous As and P during the refinement stage suggests highly selective As removal via reductive precipitation. This breakthrough separation strategy implies low levels of solid-phase impurities. Indeed, the As mass fraction of the precipitated solids was confirmed to reach >99%, with trace sulfur (S) derived from TDO typically making up the remainder (table S3). Analysis of the aqueous phase following separation of particulate As indicated that urea and several S species, including sulfite, sulfate, and thiosulfate, were the dominant aqueous products of TDO decomposition (table S4). Methods to recover P from these solutions, such as via Ca dosing to precipitate Ca-P solids, must therefore be selective for P in the presence of high levels of aqueous S species.

Overall, our results demonstrate that a high TDO/As mol ratio of 200 achieved the highest reductive precipitation efficiency (>99% conversion in some cases), but comparable performance (>85% removal) was also obtained using roughly half this TDO dose, which indicates a trade-off between reductant dose and residual aqueous As. The optimal TDO amount to implement in practice will therefore depend on system-specific priorities. For example, high TDO doses can be desirable in settings where eliminating residual aqueous As is the priority, whereas lower TDO doses may suffice when minimizing chemical use outweighs As disposal concerns. Another important practical consideration is the potential to recycle the reductant following As(0) formation. The electrochemical conversion of aqueous S-bearing reaction products into sulfoxylate, if possible under practical conditions, could markedly lower the overall chemical demand of the process, which is a benefit that warrants targeted future studies. Finally, with respect to the safety of implementing reductive As precipitation, TDO offers the notable advantage that it does not fully reduce aqueous As to arsine gas (AsH₃), the most toxic form of arsenic to humans (22). This contrasts with the common observation of AsH₃ production in side reactions using other established methods to form As(0) from aqueous As, such as chemical reduction with borohydride (37), direct electrochemical reduction (38), and photocatalyzed reduction on particle surfaces (39).

Molecular-scale structure of upcycled As(0)

Irrespective of the different initial sludge compositions, all upcycled As solids displayed nearly identical structural features (Fig. 4 and fig. S6). The position of the As K-edge x-ray absorption near edge structure (XANES) maxima for the upcycled solids matched that of commercial As(0). This result confirms that the upcycled solids consist of As in the 0 oxidation state [As(0)], which formed via reduction of As(V) contained in the initial sludge (Fig. 4). Although the XANES spectra for the upcycled and commercial As(0) solids indicated similar oxidation states, the extended x-ray absorption fine structure (EXAFS) spectra of upcycled and commercial As(0) differed substantially, consistent with different As(0) bonding environments. The EXAFS spectra of upcycled As(0), unlike commercial As(0), were characterized by smooth periodic oscillations that peaked in amplitude near 9 Å⁻¹, with little evidence for interfering waves from multiple backscattering atoms. These smooth oscillations differed from the EXAFS spectra of a suite of As standards, including commercial As(0) (fig. S7). Shell-by-shell fits of the Fourier-transformed EXAFS spectra confirmed a unique bonding environment for upcycled As(0) consisting of a single As-As path with coordination number (CN_As-As) near 4 and interatomic distance (R_As-As) of 2.46 Å (table S5 and section S1). The extremely disordered or amorphous structure suggested by a single As-As shell contrasts with the crystalline structure of commercial As(0). The fits of commercial As(0) indicated a different first-shell environment (CN_As-As = 3, R_As-As = 2.52 Å) and, most notably, required several As-As paths to fit longer-ranged atomic shells (table S5). Pair distribution function (PDF) analysis supported the different local- and intermediate-ranged structures of upcycled and commercial As(0), including the shorter first-shell As─As bond length and extreme disorder of upcycled As(0) (fig. S8 and section S2). In particular, the atomic pair correlations for upcycled As(0) decayed at R < 10 Å, consistent with a coherent scattering domain >1 nm, greatly contrasting with commercial As(0). Scanning transmission electron microscopy (STEM) images (Fig. 4 and fig. S6) indicated that disordered upcycled As(0) particles formed relatively monodisperse assemblies of 20- to 30-nm aggregates consisting of pure As with negligible trace elements, consistent with the purity measurements. Together, the characterization data for upcycled As(0) point to a largely amorphous structure that consists of subunits of fourfold coordinated As(0) that link together noncoherently to form 20- to 30-nm spherical particles (fig. S9).

Fig. 4. Structural characterization data for upcycled As(0).

(A to C) As K-edge XAS data and (D to F) TEM images. In (A), the As K-edge XANES spectra for a subset of initial sludges (red spectra) are provided to compare with upcycled As(0). (A) to (C) also include As K-edge XAS data for standards of As(III) and As(V) adsorbed to hydrous ferric oxide [As(III)_HFO, As(V)_HFO], commercial As(0), and the products of reductive As precipitation in control experiments using aqueous As(III) [Reduction Aq. As(III)], aqueous As(V) [Reduction Aq. As(V)], and synthetic sludge (Syn. sludge) as the starting materials. For these control reduction experiments, aqueous As(III) and As(V) were prepared by adding the corresponding sodium salts to ultrapure water. (D) to (F) show TEM data for the upcycled 0.9_AerFe^IIOx sample. The inset of (D) shows the selected area electron diffraction pattern.

The unique structure of upcycled As(0), which lacks intermediate- and long-ranged order, likely gives rise to material properties that differ from crystalline As(0). Amorphous metals have distinct functionality relative to their crystalline counterparts, with amorphous selenium (a-Se) applied in electronic devices serving as a pertinent example of the increased value of amorphous metal(loid) phases (40). Therefore, high-purity amorphous As(0) is an intriguing alternative to crystalline As(0) to apply directly or use as a precursor for functional materials (e.g., two-dimensional optoelectronics and semiconductors), which can further improve the economic benefits of As-rich sludge upcycling (41). The structural differences between amorphous and crystalline As(0) may have particular importance in facilitating the synthesis of arsenene, the two-dimensional form of pure As(0), which has promising semiconducting properties (17). Specifically, amorphous As(0) might be advantageous as a precursor for arsenene formation due to its lack of long-range order. Our work enables future systematic comparisons of the benefits of fabricating As-based advanced functional materials from amorphous and crystalline As(0) precursors, which is needed to best quantify the economic value of upcycled As(0).

DISCUSSION

Paradigm shift: From contaminant to commodity

The conversion of As contained in groundwater treatment sludge to form pure nanoscale As(0), all while unlocking P recovery, has the profound 2-for-1 benefit of eliminating unsustainable and hazardous sludge disposal methods while reducing the reliance on imports of CRMs derived from extractive practices. The price of metallic As(0) can vary widely depending on material properties (i.e., purity), with some high-value metallic As(0) currently sold for more than 1000 USD/kg (42), which highlights the potential economic benefits of sludge upcycling. However, focusing exclusively on the market price of As(0) overlooks the more profound environmental and societal value of valorizing As-laden groundwater treatment sludge. Notably, life cycle impact assessment (LCA) (section S3 and figs. S10 and S11) indicated the production of metallic As(0) from As-rich sludge yielded a net decrease in human health and ecotoxicity impacts compared to landfilling and conventional As(0) production from mining despite the chemical resources required for upcycling. The upcycling approach presented here can disrupt current water sector practices in all regions, but As valorization can be most societally transformative in rural, resource-scarce areas where geogenic As has for decades decimated the health and livelihood of millions. The majority of the estimated ~200 million people exposed to toxic As levels in groundwater live in marginalized communities that cannot afford conventional groundwater treatment (1, 43). A challenge for nonconventional treatment systems operating in these rural areas has been the absence of As-rich sludge disposal strategies, often leading to open disposal of this carcinogenic waste to soils and surface water (fig. S1), perpetuating the health crisis (8, 13). The valorization of As waste provides an improved sludge management solution and, through the sale of recovered materials (given increased CRM demand), can create critical economic opportunities that improve the financial viability of decentralized groundwater treatment, particularly in low-income communities. Considering that decentralized As treatment systems often fail due to prohibitive operating costs (43), coupling these systems with waste valorization for cost recovery can improve their economic sustainability. Since the efficiency of As(0) recovery increases with sludge As levels (Fig. 3), the incentive to create treatment systems that integrate As removal and upcycling is highest in the same regions plagued most by intense geogenic As contamination. Therefore, this work can catalyze the previously unimaginable scenario that the same contaminant resulting in “the largest mass poisoning in human history” (2) can be transformed to a commodity that provides increased economic welfare to local communities. Innovative treatment plants should now be designed to encompass the new role of water treatment as sources of both clean water and valuable CRMs.

MATERIALS AND METHODS

Arsenic-rich Fe oxide sludge samples

Sludge samples were obtained directly from the sludge handling and storage areas of a set of groundwater treatment plants that were selected strategically to encompass a wide range of groundwater composition (e.g., influent As and P levels) and plant properties (e.g., plant capacity). In particular, the plants used several distinct treatment processes for As removal, which leads to varying groundwater As treatment efficiencies and subsequent differences in sludge composition and structure (5). The range of treatment processes included: (i) groundwater aeration to oxidize naturally occurring Fe(II) (AerFe^IIOx), (ii) microbial-mediated oxidation of naturally occurring Fe(II) and As(III) (BioFe^IIOx), (iii) external dosage of ferric chloride solution (FeCl₃), and (iv) electrochemical anodic Fe(0) oxidation coupled with air cathode H₂O₂ production (FeElec). Treatment plants were located in the countries of Denmark, Belgium, and the United States, with additional descriptions of treatment plant characteristics provided in section S4 and table S6. Depending on sludge accessibility, 0.5 to 10 kg were collected from each treatment plant using plastic shovels, and sludge samples were sealed in plastic bags on-site before transport to the laboratory. Sludges were dried at room temperature, homogenized by hand with a mortar and pestle, and passed through a 4-mm sieve to remove large debris, such as leaves and other plant matter, consistent with standard sludge processing procedures (44). The mass fractions of major and trace elements in the dried sludges were determined by aqua regia digestions and inductively coupled plasma optical emission spectrometry (ICP-OES) analysis using a PerkinElmer Avio 550 instrument. For some sludge samples, the elemental composition was reported in our recent work (5, 30), but the composition of several additional samples is provided in the present study (table S1). We adopt a naming convention to identify the sludge samples in this work that includes the As mass fraction in gram per kilogram (total dry weight) followed by the treatment process employed by plant, which can affect sludge composition. For example, the 0.9_AerFe^IIOx sample represents sludge initially containing 0.9 g As/kg that was generated by oxidizing naturally occurring Fe(II) via aeration (table S6).

In addition to sludge samples obtained from treatment plants in the field, we prepared synthetic As-rich Fe oxide sludge in the laboratory as a reference material to compare with the field samples. The synthetic sludge was produced using analytical grade chemicals and Milli-Q water (18.2 megohm·cm). Synthetic sludge suspensions were prepared by mixing stock solutions of 500 mM FeCl₃·6H₂O (100 ml) and 10 mM Na₂HAsO₄·7H₂O (25 ml) in 200 ml of Milli-Q water. The mixture was then titrated under vigorous stirring to pH 7.5 using 1 M NaOH and brought to a final volume of 1000 ml. After 1 day of reaction, the solids were centrifuged, washed six times with Milli-Q water, and dried at room temperature. The synthetic sludge, which consists of As(V) sorbed to poorly ordered Fe(III) (oxyhydr)oxides, is termed Syn_Sludge in this work.

Upcycling of As-rich sludge

Sludge upcycling was performed using a two-stage chemical process consisting of extraction and refinement. The first step, extraction, aimed to mobilize surface-bound As from the solid sludge by addition of base to increase pH. This step involved adding dried sludge to 40 ml of 1 M NaOH solution. We selected an extraction solution consisting of 1.0 M NaOH based on our previous work investigating the type and concentration of bases to promote As and P release from Fe(III) (oxyhydr)oxides (30). The mass of sludge contained in each suspension was varied to yield several initial solid S/L ratios of 13 to 200 g/liter. The extractions were performed at room temperature in the dark and the extraction time ranged from 1 hour to 7 days. Although the use of 1 M NaOH was informed by our previous research, the current study investigates different S/L ratios and different sludge samples and includes higher temporal sampling resolution, such that the vast majority of data points reported in Fig. 2 (132 of 138) reflects extraction conditions unique to the current study. The Eh and pH values of the suspensions were recorded using a calibrated glass pH electrode (WTW SenTix@ 940) and a calibrated Pt-ring ORP electrode (Metrohm 6.0451.100), respectively. Following the reaction, the suspension was centrifuged and the supernatant was passed through 0.2-μm polyethersulfone (PES) filters. An aliquot of the solution passing the filter was removed and immediately acidified with HNO₃ for measurements of major and trace elements by ICP-OES. The remaining solution was reserved.

The second step, refinement, applied a chemical reductant to form particulate As(0) from the aqueous As extracted from sludges. In this step, 20 ml of the extraction solution was transferred to a separate 50-ml polycarbonate tube. Next, the extraction solution was heated to 80°C in a water bath, and solid TDO was added to the heated solution. At temperatures above 70°C, TDO rapidly decomposes, generating the active reducing agent, primarily SO₂²⁻, which has the ability to reduce AsO₄³⁻ to metallic As(0) under alkaline conditions (22, 33, 34). The efficiency of aqueous As removal via reductive precipitation was investigated at a range of TDO/As mol ratios (50 to 600). The reduction reaction proceeded for 1 hour, whereupon the reactor was removed from the water bath, the Eh and pH were measured after cooling to room temperature, and the suspension of dark gray particles was centrifuged. The supernatant solution was decanted and filtered through 0.2-μm PES filters, whereas the centrifuged solids were washed with N₂-purged deionized water and reserved for subsequent characterization. Several aliquots of the supernatant solution passing the filters were collected for measurements of a suite of aqueous species: Major and trace inorganic elements were quantified by ICP-OES; TDO and urea were measured by high-performance liquid chromatography (HPLC); and aqueous sulfur species (sulfate, sulfite, thiosulfate, and sulfide) were measured by ion chromatography (IC). Additional details on the HPLC and IC measurements are provided in section S5. Finally, the purity of a subset of upcycled As(0) solids was determined by repeatedly washing the solids in N₂-purged 2% H₂SO₄ and Milli-Q water, followed by complete digestion in 30% H₂O₂ and 2% HNO₃. The metals and nonmetals in the digestion were quantified by ICP-OES analysis. All extraction and upcycling experiments were replicated at least three times, with data points for aqueous chemical analyses reported as the average and SD of replicates. All structural characterization data (described below) were obtained for solids generated in one of the replicate experiments.

Characterization of upcycled As(0)

The upcycled As(0) solids were characterized by a set of advanced molecular- and nanoscale characterization techniques. Synchrotron-based As K-edge x-ray absorption spectroscopy (XAS) was used to identify the average oxidation state and local bonding environment of As in the upcycled As(0) solids and a commercial As(0) metal reference material purchased from Thermo Fisher Scientific. The As K-edge XAS data were collected at the Balder beam line of MAX IV (Lund, SE) and the P.65 beam line of DESY (Hamburg, DE). Spectra were recorded simultaneously in transmission and fluorescence modes out to k of 16.0 Å⁻¹, with data of higher quality (typically transmission measurements) reserved for subsequent analysis. All spectra were collected with samples contained in a liquid He cryostat. The maximum of the first derivative of an Au(0) foil was set to 11,919 eV to calibrate the x-ray energy. Four to 16 scans were collected for each sample, depending on data quality. Changes in line shape and peak position indicative of beam damage were examined, but no artifacts were observed. The Athena software package (45) was used to align, average, and background-subtract experimental spectra. The EXAFS spectra were extracted from the normalized XAS data using k³-weighting and were Fourier-transformed over the k-range 2 to 15 Å⁻¹ using a Kaiser-Bessel window with dk of 3 Å⁻¹.

The As K-edge EXAFS spectra of upcycled As(0) and commercial As(0) were analyzed by shell-by-shell fits. Theoretical curve fits were performed from 1 to 5.5 Å in R + ΔR-space based on algorithms derived from IFEFFIT (46). The interatomic distance (R), coordination number (CN), mean squared atomic displacement parameter (σ²), and the change in threshold energy (ΔE₀) were typically varied for each fit. Phase and amplitude functions for single scattering As─As paths were calculated using FEFF6 (46) and were derived from the crystal structure of As(0) (47). The goodness-of-fit for individual fits was assessed using the R factor, whereas the reduced chi-square (χ_ν²) parameter was used to compare multiple fits of a single spectrum when additional atomic paths were included, which was required for robust fits of the upcycled As(0) solids. Additional details of XAS data collection and the shell-by-shell fit analysis are presented in section S6.

The morphology and nanoscale aggregate structure of the upcycled As(0) particles were investigated by scanning transmission electron microscopy (STEM). For these measurements, solids from the refinement experiments were resuspended in N₂-purged Milli-Q water, and a small volume of the suspension was removed using a micropipettor and deposited on a Cu transmission electron microscopy (TEM) grid and evaporated. The TEM images were collected at the Danish Technical University NanoLab using a Titan E-Cell 80-300ST operated at 300 keV. The instrument was equipped with a Gatan OneView camera, and a standard single tile TEM holder was used. Images were analyzed using GMS3.

Structural information obtained from As K-edge XAS data was complemented by pair distribution function (PDF) analysis of high-energy x-ray scattering data for a subset of upcycled As(0) solids. Solids from the refinement experiments were dried in an anaerobic chamber and ground with a mortar and pestle. The homogenized As(0) solids were loaded in Kapton capillaries, sealed with epoxy on both ends, and stored in an anaerobic chamber until shipment to the beam line with ice packs. This procedure was followed for samples of both upcycled As(0) and commercial As(0). High-energy x-ray scattering experiments were carried out on the Wiggler High Energy beam line at the Brockhouse Diffraction Sector of the Canadian Light Source (Saskatoon, CA) (48). Data were collected out to a Q value of 26 Å⁻¹ at room temperature using 59.76-keV x-rays. The GSAS-II software package was used for data reduction and normalization and to extract the PDF [g(r)] from the total scattering data following standard protocols. Additional details regarding data collection and analysis are provided in section S6.

Life cycle impact assessment

A life cycle assessment (LCA) was performed with the specific goal to compare the human health and ecotoxicity impacts of As sludge upcycling with those of landfilling and open disposal. Because As upcycling also generates metallic As(0), the LCA also accounted for the toxicity impacts from metallic As(0) production via conventional As mining and concentrate processing. The functional unit for the LCA was 1.0 kg of metallic As(0), with all inputs normalized to this value. For example, the mass of sludge required to create 1.0 kg of As(0) via As upcycling was used as the input for the sludge disposal models to ensure accurate comparisons. The data used to create the life cycle inventory for As upcycling was derived from the extraction and refinement stages of the 0.9_AerFe^IIOx sample (S/L ratio = 200 g/liter, TDO/As mol ratio = 200) since this sample was obtained from one of the largest treatment plants in Denmark (capacity of 5,500,000 m³/year). The initial models for landfilling, open disposal, and metallic As(0) production from mining (fig. S10) were taken from our recent publication (9) but were amended such that equivalent initial mass flows were included in each system. We used the TRACI 2.1 methodology to determine human health and ecotoxicity impacts, which are expressed in comparative toxicity units (CTU/kg) (49). The OpenLCA program was used for impact calculations. Additional information on the LCA methodology and key data in the life cycle inventory are given in section S7 and tables S7 and S8.

Supplementary Materials

This PDF file includes:

Supplementary Text

Figs. S1 to S11

Tables S1 to S9

References