A landmark National Research Council report from 2013 emphasized the need to transition to passive or less resource-intensive management strategies like monitored natural attenuation (MNA) at sites where pump-and-treat and active remedies are providing diminishing returns. While this report is now more than a decade old, the process for implementing this type of transition assessment is still not well understood by practitioners. The objective of this study was to help address this gap by developing a software tool (TA2 Tool) that aids in gathering and analyzing data that are relevant for a site-specific transition assessment. The implications of site complexities on achieving remedial objectives are a key component of transition assessment. This free web-based tool has modules that perform quantitative assessment of concentration trends and project the remediation timeframe based on the current remedial approach. It has modules that predict how remediation timeframes are influenced by matrix diffusion to assess if additional remediation is warranted. Crucially, it includes modules that evaluate MNA as a transition technology, specifically by looking at plume stability, natural attenuation rates, and projections of plume concentrations at a downgradient point of compliance in the absence of further active treatment. This new tool complements existing resources on technology optimization and transitions, including ITRC guidance. The value of performing these types of assessments is highlighted through empirical data and case studies that show that it is not necessary to operate pump-and-treat systems in perpetuity and that many sites with these systems have either transitioned to other technologies or been closed.

Chlorinated ethenes are common contaminants in aerobic water-supply aquifers, and they can form large dilute plumes. Often, concentrations attenuate naturally along the flow path, and the bulk first-order rate constant for attenuation with travel time in the aquifer (k na) can be used to determine whether the concentrations at a point of compliance will be below the cleanup goal. If concentrations are below the goal, k na provides the U.S. EPA the first line of evidence for Monitored Natural Attenuation (MNA) as a remedy. However, monitoring wells may fail to sample important areas of a plume; thus, the first line of evidence alone is often not convincing for MNA acceptance. Consequently, there is a need for methods that provide additional lines of evidence to support the use of MNA in aerobic water-supply aquifers. A recently developed 14C-assay addresses this gap by directly quantifying natural abiotic degradation of chlorinated ethenes through the accumulation of radiolabeled polar products. Specifically, the assay provides rate constants for abiotic degradation within 6 weeks, offering both a quantitative measure of degradation (the second U.S. EPA line of evidence for MNA) and proof that contaminants are being transformed (the third line of evidence). Here, the 14C-assay was used to evaluate abiotic degradation at three sites. At one site k na was 0.89 +/- 0.094 per year. The rate constant for abiotic TCE degradation in a 14C-assay using sediment from the site was 0.27 +/- 0.11 per year, and the rate constant for cDCE degradation was 1.57 +/- 0.39 per year. These results demonstrate that the 14C-assay can provide robust additional lines of evidence supporting MNA in aquifers where abiotic degradation would otherwise be overlooked.

Biodegradation of petroleum hydrocarbons in groundwater occurs naturally and can be enhanced to support contaminated site management and remediation. Molecular biological tools can be used to assess the occurrence of biodegradation and monitor bioremediation efforts; however, the use of specific genes (i.e., biomarkers) to identify the rate constants of contaminant removal has not been explored for monoaromatic hydrocarbons. In this study, an approach was developed to estimate an apparent first-order rate constant for anaerobic biodegradation of toluene based on toluene concentration and the abundance of the alpha subunit of the benzylsuccinate synthase gene (i.e., bssA). The utility of the approach was evaluated by comparing the distribution of estimated rate constants to those from a published compilation at benchmark sites. There was good agreement between the distribution of rate constants calculated from the abundance of gene copies from soil cores and the distribution of rate constants at benchmark sites, while rate constants calculated from the abundance of gene copies in groundwater were lower than those from the benchmark sites. Given that groundwater samples are more common and convenient to obtain, it is proposed to use the rate constants estimated from groundwater samples to document whether the microbial community has acclimated for biodegradation of toluene. When practitioners use models to evaluate risk at petroleum release sites, they often assume a "typical" rate constant for biodegradation. If the abundance of the bssA biomarker demonstrates that the microbial community has acclimated, then it is appropriate to select and use a rate constant from benchmark sites to forecast biodegradation and attenuation of toluene at a site being evaluated.

Among the most challenging sites to remediate are those where the groundwater is contaminated with large, dilute plumes of tetrachloroethene (PCE), trichloroethene (TCE), cis-1,2-dichloroethene (cDCE), and vinyl chloride (VC). Monitored natural attenuation (MNA) may be a viable strategy, but relevant mechanisms such as abiotic degradation can be difficult to document. The objective of this study was to evaluate the use of a carbon-14 (C-14) assay to measure the rate constants for degradation of chlorinated ethenes in contaminated aquifers using samples of soil and groundwater from three sites. Use of C-14-labeled compounds makes it possible to quantify degradation by measuring the accumulation of degradation products that are otherwise difficult to discern from background levels (e.g., (CO2)-C-14 and C-14-labeled soluble compounds). The soil and groundwater samples were added to serum bottles; one set of the microcosms was incubated in the absence of oxygen, another set in the presence of oxygen. After injecting purified C-14-PCE, C-14-TCE, or C-14-cDCE, unlabeled compounds were added to bring the initial concentrations to similar to 200-1700 mu g/L. The microcosms were placed on a tumbling device to ensure gentle agitation during incubation. At weekly intervals over 42 days, 5 mL liquid samples were withdrawn, filtered, and sparged to remove the unreacted C-14-labeled parent compound. The amounts of C-14 products that accumulated were used to calculate pseudo-first-order rate constants that ranged from 0.0092 to 0.24 per year. In a companion paper by Wilson et al. (2025), the rate constants are evaluated as a line of evidence for assessing the applicability of MNA.

Microplastics (MPs) and nanoplastics (NPs) can affect microbial abundance and activity, likely by damaging cell membrane components. While their effects on anaerobic digestion are known, less is understood about their impact on microbes involved in contaminant bioremediation. Chlorinated volatile organic contaminants (CVOCs) such as tetrachloroethene (PCE) and explosives like hexahydro-1,3,5-trinitro-1,3,5-triazine (RDX) are common in the environment, and their bioremediation is a promising cleanup strategy. This study examined how polystyrene (PS) and polyamide 6 (PA6) MPs and NPs influence CVOC and RDX biodegradation. PS particles did not inhibit the CVOC-degrading community SDC-9, but PA6 MPs impaired the reductive dechlorination of trichloroethene (TCE) to cis-1,2-dichloroethene (cis-DCE), causing a “cis-DCE stall” with no further conversion to vinyl chloride (VC) or ethene. Only 45% of TCE was dechlorinated to cis-DCE, and Dehalococcoides mccartyi abundance dropped 1000-fold in 35 days with PA6 MPs. In contrast, neither PA6 nor PS MPs and NPs affected RDX biotransformation. These results highlight the significant impact of PA6 MPs on CVOC biodegradation and the need to consider plastic pollution in environmental management.

This article compares the full installation and operational costs of a hydraulic containment ("pump-and-treat," P&T) system with those of a hypothetical contemporary in situ colloidal activated carbon (CAC) barrier. Worked examples are provided using public domain data from the FT-02 fire-fighting training site on the former Wurtsmith Air Force Base in Oscoda, MI. The projected CAC costs are approximately a third of the P&T costs over projection periods of 15-100 years ($ 7.2 M vs. $ 19 M at 30 years; 38%). Hydraulic containment and CAC remediation systems prevent the spread of per- and polyfluoroalkyl substance (PFAS)-contaminated groundwater. Hydraulic containment works by extracting contaminated groundwater, removing the contamination using activated carbon or other means, and re-injecting the cleaned groundwater. The arrangement of extraction and injection wells and the related groundwater pumping rates create a hydraulic barrier that captures and contains the PFAS plume. In situ CAC barriers work as passive underground filters. Micron-scale particles of activated carbon are injected into contaminated aquifers using drilling equipment or injection wells. Once injected, the carbon particles attach to the soil. An in situ permeable barrier is installed across a contaminant plume through the injection of a number of points at suitable spacings. Groundwater flows through the barrier zone unimpeded while PFAS contamination is captured and contained by the activated carbon. The longevity of the barrier is determined by the quantity of carbon emplaced relative to the contaminant flux. Barriers are typically designed to last for years (decades), after which time, carbon re-applications can be made, if required. Principal differences between the approaches are the scale of operation and maintenance requirements, and, for P&T, the bringing of PFAS-impacted water above ground to treat. This generates a filtration medium that is contaminated with PFAS, and which therefore requires handling as a PFAS waste with attendant liability. Hydraulic containment is also an active technology (requiring external energy input) and requires the operation and maintenance of pumping and filtration equipment. In situ CAC barriers are passive (powered by natural groundwater flow) and have no operation and maintenance requirements. They do not bring PFAS-impacted material above ground and do not generate waste. Performance data of the installed P&T hydraulic containment system were analyzed to estimate the time to remedial completion using the system alone. Data extrapolation supported by statistical analysis indicates clean-up targets will not be reached within 100 years of pumping. It is not realistic for P&T to be regarded as a means of aquifer clean-up as the aquifer will remain contaminated for the realistic future. Comparison is made between P&T and CAC on their common basis as containment approaches. The goal is to reduce the exposure of down-gradient receptors to PFAS.

The intricate interplay between climate and tectonics profoundly shapes landscapes over time frames surpassing 10 million years. Active tectonic processes and climatic shifts unsettle established drainage systems, instigating fragmentation or amalgamation of watersheds. These activities yield substantial transformation in surface hydrologic connectivity, thereby underlining the profound influence of these tectonic and climatic forces on the evolution of both landscape and hydrology. Such transformations within the hydrological landscape have direct implications for the evolution of aquatic species. As connections among aquatic habitats undergo reconfiguration, they incite shifts in species distribution and adaptive responses. These findings underscore the role of tectonics and climate in not only sculpting the physical landscape but also steering the course of biological evolution within these dynamically changing aquatic ecosystems relying on hydrologic connections. Despite the significance of these interactions, scholarly literature seldom examines alterations in hydrologic connectivity over tectonic, or orogen-scale, timescales. This study aims to bridge this gap, exploring changes in hydrologic connectivity over extended periods by simulating a continental rift system akin to the Rio Grande Rift, USA, subject to various tectonoclimatic scenarios. Multiple rift basins hosting large lakes, brought into existence by active tectonic extension, are further molded by tectonic extension and post-rift climatic changes. The study focuses on phenomena such as interbasin river breakthroughs and knickpoint generation, assembling a time-series of connectivity metrics based on stream network characteristics such as flow rate, flow distance, and captured drainage areas. We anticipate that the insights gleaned from this study will enhance our comprehension of the enduring impact of tectonic and climate processes on hydrologic connectivity and the subsequent evolution of aquatic species.

When perfluoroalkyl and polyfluoroalkyl substance (PFAS)-contaminated groundwater is a source—or potential source—of drinking water, remediation strategies are commonly selected with a goal of protecting the health of human populations who may use the contaminated water for extended periods. Long-term human exposure to certain PFAS at environmentally relevant levels has been associated with increased serum cholesterol, decreased vaccine response, and an increased risk of cancer (Interstate Technology and Regulatory Council [ITRC], 2023; US Environmental Protection Agency [US EPA], 2023a, 2023b). Recently, the International Agency for Research on Cancer (IARC) classified perfluorooctanoic acid (PFOA) as "carcinogenic to humans" (IARC Group 1) and perfluorooctane sulfonic acid (PFOS) as "possibly carcinogenic to humans" (IARC Group 2B) (Zahm et al., 2023). Remediation strategies are chosen that can decrease PFAS levels in water to applicable health-based criteria and thereby limit exposure of local populations to PFAS through ingestion of drinking water. However, this approach does not consider the potential for human exposure throughout the life cycle of the remediation technology, in which spent media may need to be disposed of, regenerated, or destroyed over the many years the technology is likely to be in place. In this commentary, we consider four PFAS remediation technologies and identify those places in the life cycle that have the potential for environmental releases from the handling, transport, disposal, regeneration, and/or destruction of remediation wastes. Importantly, we also identify where those releases have the potential to result in human exposure to PFAS, focusing on the long-chain perfluoroalkyl acids (PFAAs) and using PFOA and PFOS as examples. Technologies that have been demonstrated at multiple sites, under diverse conditions, by multiple practitioners, are commercially available, and are well documented in practice or peer-reviewed literature. Field-implemented technologies have been demonstrated to meet site-specific PFAS treatment objectives, at the intended final application scale, and are widely accepted in the regulatory and scientific community. Four technologies are listed within this category for treatment of PFAS-impacted liquids that have relevance to groundwater/extracted groundwater. The technologies are (1) granular activated carbon (GAC), (2) ion-exchange resin (IX), (3) foam fractionation (FF), and (4) in situ remediation with colloidal activated carbon (CAC). It is these technologies that the present commentary considers. It is noted that a category of "high-pressure membranes" is listed within the ITRC "field-implemented" technology category for liquid PFAS treatment and as a proposed treatment technology by US EPA (2023c). However, this technology is omitted from the present commentary, as its application is principally in point-of-use drinking water purification systems—often as part of a treatment train—rather than in groundwater remediation. Analogous life cycle considerations for wastes generated by this technology would nevertheless also apply. Importantly, three of the considered remediation technologies are variations of "pump and treat" and produce waste over the entire life of the project. That waste must first be managed in the field and then transported for disposal, regeneration, or destruction. The management of these remediation wastes has the potential to release PFAS into air, soil, surface water, and/or groundwater, with the attendant potential for human exposure at many points in the waste cycle. In contrast, passive in situ remediation with CAC—the fourth technology—generates no waste. Pump and treat remediates a contaminated aquifer by extracting the contaminated groundwater; this water is then treated before being returned to the aquifer or discharged to surface water. Pump and treat has been used extensively for groundwater remediation for over 40 years (National Research Council [NRC], 2013). Most of this experience has been for chlorinated solvents such as trichloroethene (TCE) (NRC, 2013). Pump and treat is not particularly effective as a remedy for chlorinated solvents at many sites because the solvents tend to sorb to aquifer materials or diffuse into nontransmissive portions of the aquifer where they are less available to be extracted (Chapman & Parker, 2005; Guo et al., 2019; Mackay & Cherry, 1989; NRC, 1994). The same situation applies to PFAS compounds. As discussed below, pump and treat should be even less effective for long-chain PFAS compounds such as PFOA or PFOS. PFOA or PFOS sorb more strongly to aquifer solids than, for example, TCE. The potential for sorption in aquifers is typically attributed to organic carbon in the aquifer solids. The effects of sorption can be predicted from the partition coefficient between aquifer water and organic material in aquifer solids [Koc (L/kg)]. The Koc for PFOA is approximately 448 L/kg (geometric mean of 37 values reported in tab. 4-1 of ITRC, 2023). The Koc of PFOS is approximately 2380 L/kg (geometric mean of 43 values reported in tab. 4-1 of ITRC, 2023). In contrast, the Koc of TCE is approximately 94 L/kg (geometric mean of 21 values reported in US EPA, 1996). Therefore, based on a comparison of Kocs, PFOS sorbs to aquifer solids 20 times more strongly than TCE, and PFOA sorbs five times more strongly. A much smaller proportion of the PFOS or PFOA in the aquifer will therefore be in the groundwater and available for extraction by pump and treat. Consequently, more pumping will be required to remediate the aquifer. The significance of the strength of aquifer absorption in field–implementation, is a slowing of the removal rate of PFAS contamination and the consequent extension of the remedial timeline. This is driven by the sorbed mass effectively acting as a "reservoir" from which groundwater is recontaminated through desorption to establish a new equilibrium as remediation progresses. This process will continue, slowing the decline in groundwater concentration, until the reservoir itself is depleted. Taking PFOS as an example and using a soil density of 1.7 g/cm3, a porosity of 0.23, an organic carbon fraction of 0.2%, and the above Koc value, the sorbed mass reservoir would be 35 times the aqueous mass at equilibrium—a "reservoir factor" of 35. This contrasts with an equivalent reservoir factor of 1.4 for TCE under the same aquifer conditions. The chlorinated solvents are also removed from groundwater through destructive natural attenuation processes such as biodegradation and abiotic degradation (Lebrón et al., 2015; Wiedemeier et al., 1998). These degradation processes contribute to remediation and reduce the time needed to attain a remedy. However, organisms that can metabolize perfluorinated organic compounds are rare (Wackett, 2021). Studies compiled by the PFAS Team of the ITRC show that natural degradation in groundwater can transform precursors to form PFOA and PFOS, but the team does not cite any studies that claim significant degradation of PFOA or PFOS in aquifer materials (ITRC, 2023). It is unlikely that natural degradation processes will contribute to pump and treat remedies for recalcitrant perfluoroalkyl substances, such as PFOA or PFOS. Considering the greater propensity of PFOA and PFOS to partition onto aquifer materials compared to chlorinated solvents and the unlikely contribution of degradation mechanisms to removal of these contaminants, it follows that pump and treat will require significantly longer periods of time to achieve a desired reduction in concentrations of PFOA and PFOS compared to chlorinated solvents. Given that pump and treat clean up of chlorinated solvents at many sites may require many decades (Mackay & Cherry, 1989; NRC, 1994, 2013; Travis & Doty, 1990), it is unrealistic to consider the approach as a means of cleaning an aquifer. Rather, pump and treat provides an effective strategy for ongoing plume capture and containment. PFAS contamination must be removed from groundwater extracted by a pump and treat system before the water may be returned to the aquifer or discharged. Contemporary treatment commonly employs sorption media for this purpose (GAC, IX). PFAS mass is transferred from the water to the sorption media, which then requires treatment or disposal. PFAS may also be separated from extracted water using FF. All these processes yield concentrated PFAS wastes. The handling, transport, and eventual disposal or destruction of the accumulated PFAS mass required in each case are components of the remediation life cycle. These approaches for extracted water treatment—GAC, IX, and FF—represent three of the field-implemented PFAS treatment technologies (ITRC, 2023) that are considered in this commentary. These technologies and their associated waste life cycles, plus the fourth technology of in situ remediation with CAC, are discussed in the following sections. GAC is widely applied as a sorption medium for separating PFAS from water (ITRC, 2023). Extracted groundwater is moved through packed beds of activated carbon; PFASs adsorb to the surface of the carbon and are removed from the water without degradation (ITRC, 2023). GAC removes PFOA, PFOS, PFHxS, and other PFAS (McCleaf et al., 2017), with perfluorosulfonic acids (PFSAs) generally removed more readily than perfluorocarboxylic acids (PFCAs) and long-chain PFAAs removed more efficiently than short-chain PFAAs (Appleman et al., 2014; McCleaf et al., 2017). When the surface area of the GAC can no longer effectively adsorb additional PFAS, the spent GAC can either be treated by incineration, disposed of in a landfill, or thermally regenerated for reuse (US EPA, 2020; Wang et al., 2022; Watanabe et al., 2016; Xiao et al., 2020). Depending on the equipment and operating conditions, both regeneration and incineration may release gas-phase PFOA, PFOS (Watanabe et al., 2016), and other PFAS to air if temperatures less than 1000°C/1800°F are used (US EPA, 2020; Wang et al., 2022; Watanabe et al., 2016). Both thermal treatment and incineration may also produce ash, which requires landfill disposal, an activity that holds the potential for release of PFAS to air and other environmental media (US EPA, 2020; see also Figure 1 and Table 1) should the incineration or regeneration not be complete (DiStefano et al., 2022). If the spent GAC or ash is placed in a lined landfill, PFAS can desorb into the landfill leachate (US EPA, 2020). PFASs were present in ash from a municipal solid waste incinerator (Liu et al., 2021) and in landfill leachate from incinerator ash monofills (Liu et al., 2022; Solo-Gabiele et al., 2020). PFAAs have been regularly detected in landfill leachate, with PFOA being one of the most frequently detected PFCAs (Hamid et al., 2018). Lang et al. (2017) detected PFOA, PFOS, and PFHxS in more than 50% of US landfill leachate samples. Air, soil, surface water, groundwater, drinking water Inhalation, ingestion, dermal contacta Air, soil, surface water, groundwater, drinking water Inhalation, ingestion, dermal contacta Air, soil, surface water, groundwater, drinking water Inhalation, ingestion, dermal contacta Landfill leachate is often disposed of to a wastewater treatment plant (WWTP), where treated water may be discharged to holding ponds or directly to surface water. Because conventional municipal wastewater treatment does not degrade or consistently remove PFAS (Schultz et al., 2006; Sinclair & Kannan, 2006), PFASs originally present in the landfill leachate will re-enter the environment after treatment, where they may migrate to soil, surface water, or groundwater. PFAS present in wastewater may also sorb to biosolids (Arvaniti & Stasinakis, 2015); if land-applied for agriculture, biosolids may also be a source of PFAS to soil, surface water, or groundwater (Johnson, 2022), although Pepper et al. (2021) found limited migration of PFASs in biosolid-impacted soils. If GAC is disposed of in an unlined landfill (leachate is not collected), PFAS, if desorbed, may migrate to soil, surface water, or groundwater in processes analogous to those described (above) for treated wastewater (ITRC, 2023). Landfills also may release PFAS to air, with PFOA and PFOS detected in the gas phase of ambient landfill air (Ahrens et al., 2011), and ionic PFOA, PFOS, and other PFAAs detected in all samples of particulate-phase landfill air (Weinberg et al., 2011). PFASs that are released to air may be subject to long-range transport (Faust, 2023). Once contaminated with PFAS, air and other environmental media hold the potential for human exposure via inhalation (air) or by incidental ingestion or dermal contact with soil or water. The handling and transport of PFAS wastes may also result in environmental releases and human exposure over the course of the remediation life cycle. While waste handling will likely be conducted under standard operating procedures designed to protect workers and the environment, both waste handling and the transport of waste to disposal facilities hold the potential for accidental releases of PFAS. The biological significance of any PFAS exposure will depend on human contact with each environmental medium, the frequency and length of that contact, on PFAS concentrations, and on the toxicological characteristics of the PFAS (see Figure 1 and Table 1). Ex situ IX resins remove PFAS from pumped groundwater based on both adsorption and on the ionic interaction of negatively charged PFAS with positively charged sites on the IX resin (McCleaf et al., 2017; Woodward et al., 2017). PFASs are not degraded by IX treatment. Both single-use and regenerable IX resins are available for PFAS removal, with single-use IX generally more effective at removing PFSAs than PFCAs (ITRC, 2023). Spent single-use resins are disposed of either by incineration or in a landfill (ITRC, 2023). Saturated IX resins can be regenerated using either an organic solvent, inorganic salts, or both, with the selected regeneration process dependent on resin properties and PFAS functional groups, among others (Dixit et al., 2021). IX regeneration yields PFAS-rich brines whose disposal currently requires incineration (Dixit et al., 2021) or landfilling (ITRC, 2023) and associated handling and transport. Other emerging disposal approaches show promise but are less well-documented (ITRC, 2023). Examples include supercritical water oxidation (Sahle-Demessie et al., 2022) and electrochemical oxidation (US EPA, 2021). Irrespective of potential destruction efficiency, these share waste handling and transport considerations common to established treatment approaches. Thus, both single-use IX resins and IX regeneration may release PFAS to air, soil, surface water, and groundwater via the mechanisms and processes discussed for GAC. Once PFAS are released, there is the potential for human exposure via inhalation, incidental ingestion, or dermal contact (Figure 1 and Table 1). Ex situ FF is a physical and chemical process in which amphiphilic PFASs such as long-chain PFAAs are separated from pumped groundwater by introducing air bubbles that rise through a column of contaminated water. PFAS preferentially accumulate at the air–water interface and therefore effectively sorb to the surface of the bubbles, accumulating in a PFAS-enriched foam at the top of the water column (Buckley et al., 2021). PFOA, PFOS, and PFHxS were efficiently removed from water in field trials (Burns et al., 2021, 2022), landfill leachate (Wang et al., 2023), and in a laboratory-scale system (Smith et al., 2022). The concentrated foam waste that is a by-product of FF can be disposed of by incineration or by landfilling (Burns et al., 2021; Wang et al., 2023) subsequent to handling and transport of the wastes. In addition to potential releases associated with incineration or landfilling of FF waste (see GAC discussion), high levels of PFOA, PFOS, PFHxS, and other PFAAs were detected in air (gas phase) and in aerosols of a pilot FF system (Smith et al., 2022). Although such releases may be effectively mitigated through suitable control measures, inhalation exposure remains of potential relevance. Thus, FF may pose additional human health concerns beyond those associated with spent GAC or IX disposal, regeneration, or destruction, in that releases of PFAS in the immediate vicinity of the FF system could expose worker populations to airborne PFAS unless appropriate and functional air pollution controls (APCs) are in place. However, even if APCs are used, they represent another waste stream that requires disposal or destruction over the life of the FF system. Injectable in situ CAC offers a means of engineering the capacity of aquifer solids to sorb PFAS. Equilibrium sorption of PFAS to CAC may be quantitatively described with isotherms (Carey et al., 2022). These provide a mathematical description of the equilibrium between sorbed and dissolved phases at different concentrations of contaminant and quantities of CAC. If the PFAS species isotherm and the quantity of applied CAC are known, the equilibrium sorption concentration for a given quantity of the PFAS species can be calculated. From this, the quantity of CAC necessary to attain the regulatory goal for PFAS remaining in the groundwater can be determined. A CAC application may typically deposit 0.1%–0.5% CAC in the target zones as a proportion of soil mass. Measured field averages of 0.02%–0.76% have been reported by Carey et al. (2022). Since the sorption of PFAS to CAC compared to natural soil organic carbon is very high, the CAC emplacement is sufficient to locally increase the coefficient of distribution (Kd) of PFAS in soil by two to three orders of magnitude at most sites. Groundwater PFAS concentrations are consequently reduced by a corresponding degree. The increased sorption of PFAS slows plume migration through proportionally increasing the contaminant retardation factor. The net result is comparable by analogy to changing the plume migration rate in sand to the rate in clay. The difference would be that, in the case of CAC addition, the contaminant migration is slowed through retardation rather than through a reduction in the velocity of groundwater. Contaminant migration is therefore slowed significantly while groundwater flow remains unaffected. Remediation is consequently achieved through containing the plume and reducing/eliminating downgradient exposure to migrating PFAS. Remedial applications may take the form of in situ permeable sorptive barriers (Carey et al., 2024). These passively treat groundwater as it flows through the treatment zone. Each 100 ft (30 m) length of CAC barrier installed through a 20 ft (6 m) depth will filter approximately 6000 gallons (23,000 L) of water per year for every 100 ft (30 m) per year of seepage velocity. For example, this would equate to approximately 1.5 million gallons (300 million liters) of water over a 20-year service life of a 500 ft (150 m) barrier at 240 ft/year (73 m/year), with no pumping required. The sorption of PFAS onto emplaced CAC restricts further migration of a PFAS plume. This may contain contamination for years, reducing exposure to populations who rely on the aquifer for drinking water (Carey et al., 2024). CAC has effectively decreased aquifer concentrations of PFOA and PFOS as well as a number of other PFAS as monitored over a 5-year period (Carey et al., 2022). Carey et al. (2019, 2024) modeled the longevity of CAC at a PFAS project site and determined that CAC could be effective for decades. Longevity of the CAC may be enhanced by increasing the amount of CAC used initially or by additional CAC injections (Carey et al., 2019, 2024). Periodic reapplications of CAC to extend the functional life of a PFAS barrier may be undertaken under a planned maintenance program, analogous to hardware and infrastructure servicing and replacement on a pump and treat containment system. A planned service period of circa 20 years would be realistic in either case (Carey et al., 2019; Department of Energy, 2001; Tutterow et al., 2002). Because CAC remains in the aquifer sorbed to solids, there is no waste management (i.e., disposal, regeneration, or destruction) and no human exposure from the use of this remediation technology (Table 1 and Figure 1). Throughout decades of manufacturing and use, numerous and diverse sources have released PFOA, PFOS, and other PFAS to the environment where they have been detected in air, sediment, soil, water, wildlife, and humans (Agency for Toxic Substances and Disease Registry, 2021; ITRC, 2023; US EPA, 2023a, 2023b). Remediation of PFAS in groundwater is a relatively recent activity, and the disposal or destruction of PFAS-contaminated remediation wastes may not have contributed substantially to previous detections of PFAS in humans or the environment. However, considering the widespread distribution of PFAS-contaminated groundwater documented by US EPA's Third Unregulated Contaminant Monitoring Rule (UCMR3) sampling (Hu et al., 2016) and the ongoing nation-wide drinking water investigations of PFAS in UCMR5 (US EPA, 2023d), PFAS remediation activities will likely increase substantially in frequency and scale. Pump and treat may contain and reduce PFAS mass within a plume or aquifer but will not recover all of the contamination in a practical amount of time. The associated handing and transport of wastes will lead to increased opportunities for accidental releases, and the disposal, regeneration, or destruction of an increasing amount of PFAS-contaminated remediation waste has the potential to contribute to human exposure to PFAS released to air, soil, surface water, and groundwater for many years. The persistence of PFAAs and certain other PFAS (Organization for Economic Cooperation and Development, 2013; Wang et al., 2015) coupled with the potential to bioaccumulate (Burkhard, 2021; Lesmeister et al., 2021) and biomagnify (Houde et al., 2011; Kannan et al., 2005; Munoz et al., 2022) increases the significance to humans of environmental releases. Releases of PFAS to the environment can be mitigated—and human exposure reduced—through the use of APCs, remediation of landfill leachate, and other methods. However, exposure to PFAS in remediation waste can be prevented by the use of a remediation technology that does not generate waste in the first place. CAC is an example of a technology that is capable of reducing PFAS concentrations in groundwater without contributing to the redistribution of PFAS in the environment or to potential long-term human exposure associated with this redistribution. We thank and acknowledge F. Javier Chalini for the preparation of the figure, which accompanies this Commentary (instagram.com/chaliniart). Funding was provided by Regenesis, San Clemente, CA, USA. Linda C. Hall, BA, MA, PhD, is an independent consulting toxicologist whose practice has focused on PFAS since 2016. She is in her eighth year as co-leader of the Regulations, Toxicity and Risk Assessment Group of the ITRC PFAS Team. In that role, she serves as an editor, contributing author, and coordinator of Fact Sheets and chapters of the PFAS Technical Regulatory Guidance document. She has nearly 40 years of experience in evaluating human exposure to and the toxicology of environmental contaminants. John T. Wilson, BS, MA, PhD, is the principal scientist at Scissortail Environmental Solutions, LLC. For 35 years he was a research microbiologist at US EPA, working on biodegradation and natural attenuation of chlorinated hydrocarbons and fuel components in groundwater. Jeremy G. A. Birnstingl, BSc, PhD, FRSC, CEnv, is Vice President—Environmental Technology at Regenesis where he is responsible for new technology establishment and the design and support of remediation projects including modeling and design-software development. He has 35 years of experience in environmental remediation specializing in groundwater cleanup and aquifer restoration. Data sharing is not applicable to this article as no new data were created or analyzed in this study.

When a remedy is selected for groundwater contamination, the monitoring record is often evaluated to extract rate constants for attenuation of contaminants over time, and the rate constants are used to forecast a time when the concentrations will attain a clean-up goal. These evaluations typically assume the rate constant does not change. Data from 11 monitoring wells at the former Twin Cities Army Ammunition Plant (TCAAP) were evaluated to test whether this assumption was valid for this site. A previous evaluation at this site (based on data from 1987 through 1999) extracted rate constants that would bring the concentrations of trichloroethylene (TCE) in the 11 wells to the maximum contaminant level (MCL) on or before 2013. By 2020, only four wells had reached the MCL. Piecewise linear regressions were used to model the relationship between time elapsed and concentration of TCE for each well, and to identify any changepoints in the rate of attenuation over time. Each well had at least one changepoint with different TCE attenuation rates on either side of the changepoints. The slope of the most recent segment after the last changepoint provides the best information to forecast concentrations in the future. For four of the wells, that forecast indicated that concentrations of TCE would never reach the MCL. Piecewise linear regression analysis proved to be a valuable tool to detect changes in rate constants, and to update forecasts of the time required for groundwater concentrations to reach a cleanup goal. Analysis of groundwater TCE concentration data collected since 1987 reveals distinct, statistically significant changepoints in the rates of TCE attenuation. Piecewise regression analysis is a valuable tool in detecting changes in rates of groundwater contaminant attenuation.image

We develop a robust and simple rule-based algorithm to autonomously simulate alluvial fan deposition and evolution under continuously developing landscape conditions without prescribing deposition locations or imposing topographic constraints. Augmented with this algorithm, landscape evolution models are capable of dynamically detecting locations of potential fan deposition by statistical measures of surface topography and fluvial dynamics, then depositing fan sediments where and when the developed conditions require. To assess the method's efficacy in depositing sediment at a mountain-valley transition zone characterized by a transport surface that permits unobstructed exit of sediment and water, a hypothetical scenario is created that involves a frontal, normal fault. It is followed by a series of sensitivity analyses to ascertain the influence of parameters affecting fan deposition and secondary processes. Uplift (u) and precipitation significantly impact fan morphological characteristics, which are within the range of real-world fans. Higher rates of each cause the notable expansion of the fan area except in cases of exceptionally high precipitation rates. Fan area has a power-law relationship with most of the tested parameters, A(f) alpha u(0.94) mu(0.31) l(f)(-0.14) K(max)(-0.65)A(c)(beta), where mu is erodibility (lithology), l(f) and K-max are fluvial parameters, and A(c) is catchment area (beta similar to 0.9). This study is the first showcasing fan power-law relationships using numerical modelling. While fan area increases with precipitation, there exists a threshold beyond which fan area diminishes, and the formation of fans ceases altogether. The algorithm provides a basis for improving mechanistic understanding of fans by offering a robust platform for testing process dominance and scaling. The results demonstrate its applicability for landscape evolution simulation over a long time and broad spatial scales. We also investigate the hydrological significance of including autonomously generated alluvial fans in coupled landscape evolution-hydrology models that focus on groundwater as well as surface water hydrology.

Chlorinated volatile organic compound (cVOC) degradation rate constants are crucial information for site management. Conventional approaches generate rate estimates from the monitoring and modeling of cVOC concentrations. This requires time series data collected along the flow path of the plume. The estimates of rate constants are often plagued by confounding issues, making predictions cumbersome and unreliable. Laboratory data suggest that targeted quantitative analysis of Dehalococcoides mccartyi (Dhc) biomarker genes (qPCR) and proteins (qProt) can be directly correlated with reductive dechlorination activity. To assess the potential of qPCR and qProt measurements to predict rates, we collected data from cVOC-contaminated aquifers. At the benchmark study site, the rate constant for degradation of cis-dichloroethene (cDCE) extracted from monitoring data was 11.0 ± 3.4 yr-1, and the rate constant predicted from the abundance of TceA peptides was 6.9 yr-1. The rate constant for degradation of vinyl chloride (VC) from monitoring data was 8.4 ± 5.7 yr-1, and the rate constant predicted from the abundance of TceA peptides was 5.2 yr-1. At the other study sites, the rate constants for cDCE degradation predicted from qPCR and qProt measurements agreed within a factor of 4. Under the right circumstances, qPCR and qProt measurements can be useful to rapidly predict rates of cDCE and VC biodegradation, providing a major advance in effective site management.

Monitored Natural Attenuation (MNA) is a preferred remedy for sites contaminated with 1,4-dioxane due to its low cost and limited environmental impacts compared to active remediation. Having a robust estimate of the rate at which biodegradation occurs is an essential component of assessing MNA. In this study, an assay was developed using 14C-labeled 1,4-dioxane to measure rate constants for biodegradation based on accumulation of 14C products. Purification of the 14C-1,4-dioxane stock solution lowered the level of 14C impurities to below 1% of the total 14C activity. This enabled determination of rate constants in groundwater as low as 0.0021 yr-1, equating to a half-life greater than 300 years. Of the 54 groundwater samples collected from 10 sites in the US, statistically significant rate constants were determined with the 14C assay for 24. The median rate constant was 0.0138 yr-1 (half-life = 50 yr); the maximum rate constant was 0.367 yr-1 (half-life = 1.9 yr). The results confirmed that biodegradation of 1,4-dioxane is occurring at 9 of the 10 sites sampled, albeit with considerable variability in the level of activity. The specificity of the assay was confirmed using acetylene and the absence of oxygen to inhibit monooxygenases.

Options for remediating 1,4-dioxane at groundwater sites are limited due to the physical-chemical properties of this compound. The relevance of natural attenuation processes for 1,4-dioxane was investigated through data from field, lab, and modeling efforts. The objectives were to use multiple lines of evidence for 1,4-dioxane biodegradation to understand the prevalence of this activity and evaluate convergence between lines of evidence. A C-14-1,4-dioxane assay confirmed 1,4-dioxane biodegradation at 9 of 10 sites (median rate constant of 0.0105 yr(-1) across wells). Site-wide rate constants were established using a calibrated fate and transport model at 8 sites (median = 0.075 yr(-1)). The C-14 assay constants are likely more conservative, and variability in rates suggested that biodegradation at sites may be localized. Stable isotope fractionation was observed at 7 of 10 sites and served as another direct line of evidence of in situ biodegradation of 1,4-dioxane. This includes sites where indirect lines of evidence, including geochemical conditions or genetic biomarkers for degradation, would not necessarily have been supportive. This highlights the importance of collecting multiple lines of evidence to document 1,4-dioxane natural attenuation, and the widespread prevalence of biodegradation suggests that this process should be part of long-term management decisions.

Several regulatory agencies recommend screening petroleum vapor intrusion (PVI) sites based on vertical screening distance between a petroleum hydrocarbon source in soil or groundwater and a building foundation. U.S. Environmental Protection Agency (U.S. EPA) indicate the risk of PVI is minimal at buildings that are separated by more than 6 feet (1.8 m) from a dissolved-phase source and 15 feet (4.6 m) from a light nonaqueous phase liquid (LNAPL) source. This vertical screening distance method is not, however, recommended at sites with leaded gasoline sources containing ethylene dibromide (EDB) because of a lack of field data to document EDB attenuation in the vadose zone. To help address this gap, depth-discrete soil-gas samples were collected at a leaded gasoline release site in Sobieski, Minnesota (USA). The maximum concentration of EDB in groundwater (175 mu g/L) at the site was high relative to those observed at other leaded gasoline release sites. Soil gas was analyzed for EDB using a modification of U.S. EPA Method TO-14A that achieved analytical detection limits below the U.S. EPA Vapor Intrusion Screening Level (VISL) for EDB based on a 10(-6) cancer risk (<0.16 mu g/m(3)). Concentrations of EDB in soil gas above LNAPL reached as high as 960 mu g/m(3) and decreased below the VISL within a source-separation distance of 7 feet. This result coupled with BioVapor model predictions of EDB concentrations indicate that vertical screening distances recommended by regulatory agencies at PVI sites are generally applicable for EDB over the range of anticipated source concentrations and soil types at most sites.

Field and numerical studies suggest that baseflow is composed of waters from a spectrum of groundwater flow paths termed the Groundwater Flow System (GWFS) - from shallow hillslope contributions to watershed-scale deep circulation originating in headwaters and discharging into lowland rivers. Here, we explore the evolution of the GWFS under prolonged droughts to understand its dynamics and multiscale nature, and to elucidate its role in baseflow generation and recession at the watershed scale. We consider three drought scenarios of varying severity and simulate groundwater flow in a 2-D cross-section of an idealized watershed with deep permeable bedrock, tracking the evolution of flow paths, baseflow, and residence times during the recession process. We find that baseflow generation at different drainage stages, and within different subwatersheds, is influenced distinctly by flow paths of different scales, depending on the relative strength of the flow paths and the position of the subwatersheds relative to the recharge/discharge zones of the deeper watershed-scale groundwater circulation. Despite having the same local relief, geology, and climate, baseflow from each sub watershed has a distinct recession behavior and time-dependent residence time distribution. Also, the hydraulic and transport characteristics of baseflow generation co-evolve and are strongly affected by the connection state of the water table to subwatersheds. These findings suggest that asynchrony and dissimilarity of baseflow generation from hillslopes under the impact of the watershed-scale groundwater flow, and interactions with local-scale and intermediate-scale groundwater flow, must be taken into account when interpreting baseflow recession data and building conceptual baseflow models at the watershed scale.

A series of quantitative decision frameworks were developed to aid in the evaluation of Monitored Natural Attenuation (MNA) as a remedy for 1,4-dioxane, chlorinated ethanes, and chlorinated ethenes. The end product is an updated version of BioPIC, a successful, existing decision framework for MNA of chlorinated ethenes. In evaluating MNA, the goal is to follow existing protocols but also incorporate more recent insights that have increased the potential viability of this technology for a wider number of contaminants. This includes data supporting the first line of evidence for MNA (decreasing concentration trends over time at appropriate sampling locations), as well as data that will serve as secondary and tertiary lines of evidence for MNA. For the BioPIC update, a similar lines of evidence approach was developed, using information from literature and project-specific data to identify the relevant lines of evidence. As part of the update, a model for predicting contaminant trends over time and distance, including a method to estimate site-specific biodegradation rate constants for chlorinated ethenes, chlorinated ethanes, and 1,4-dioxane, has been included within BioPIC. The quantitative software tool provides step-by-step technical guidance for collecting and evaluating other lines of evidence, including data associated with targeted 14 C assays, isotope fractionation, biomarkers for degradation, geochemical parameters, and co-contaminant concentrations, to improve the MNA evaluation process.

A series of quantitative decision frameworks were developed to aid in the evaluation of Monitored Natural Attenuation (MNA) as a remedy for 1,4-dioxane, chlorinated ethanes, and chlorinated ethenes. The end product is an updated version of BioPIC, a successful, existing decision framework for MNA of chlorinated ethenes. In evaluating MNA, the goal is to follow existing protocols but also incorporate more recent insights that have increased the potential viability of this technology for a wider number of contaminants. This includes data supporting the first line of evidence for MNA (decreasing concentration trends over time at appropriate sampling locations), as well as data that will serve as secondary and tertiary lines of evidence for MNA. For the BioPIC update, a similar lines of evidence approach was developed, using information from literature and project-specific data to identify the relevant lines of evidence. As part of the update, a model for predicting contaminant trends over time and distance, including a method to estimate site-specific biodegradation rate constants for chlorinated ethenes, chlorinated ethanes, and 1,4-dioxane, has been included within BioPIC. The quantitative software tool provides step-by-step technical guidance for collecting and evaluating other lines of evidence, including data associated with targeted C-14 assays, isotope fractionation, biomarkers for degradation, geochemical parameters, and co-contaminant concentrations, to improve the MNA evaluation process.