A Half-Century Perspective on the Microbial Bioremediation of PCBs

Abstract

Microbial biotransformation of PCBs was first observed in the field and in laboratory studies nearly five decades ago. Following years of investigation (on both aerobic biodegradation or mineralization and anaerobic dechlorination), bioremediation technologies are emerging for PCB-contaminated sediments. This perspective examines the arc of scientific and engineering developments that eventually brought us to today’s approach of combining anaerobic and aerobic metabolism in an in situ bioaugmentation application. Although an understanding of aerobic biodegradation of lesser chlorinated biphenyls was developed early on, more extensively chlorinated congeners were found to be resistant to oxidative transformation. Major discoveries in the microbiology of anaerobic PCB dechlorination remained necessary, including understanding of pathways, PCB-dechlorinating bacteria, kinetics, and the interaction of biological processes with the physical environment of sediment-associated PCBs. Recent developments linking microbial kinetics to dissolved-phase PCBs and identification of factors that limited bacterial growth on PCBs are eliminating bottlenecks and have facilitated field translation. A focus in the last decade on scaling the growth of a PCB-dechlorinating bacterial species to a commercially viable level, and developments in field application methods for in situ bioaugmentation have facilitated multiple field demonstrations of PCB bioremediation. These field trials, in synergy with in situ remediation of sediments with activated carbon amendment, provide a path to achieve rapid reduction of PCB exposure to the ecosystem while significantly enhancing the rate of biodegradation.

Introduction

The impetus for research on polychlorinated biphenyls (PCBs) in the 1970s originated from the discovery of the widespread nature of sediment contamination, persistence in the environment, strong potential for bioaccumulation, and alarming findings of toxicity to a wide range of animals. Early toxicology research on the pharmacodynamics of PCBs in laboratory animals reported PCB dechlorination, which was attributed to mammalian cellular metabolism but possibly involved dechlorination by gut microbiota. Microbial dechlorination of PCBs has been reported in numerous studies in the last several decades, but these recalcitrant pollutants have persisted in the environment and continue to impact human health and ecosystems. The underlying rate limitations of PCB microbial dechlorination were not understood well enough to be able to scale up and engineer the process in the field. As a result, ironically, the site where microbial dechlorination and degradation were first documented by scientists nearly 40 years ago (Hudson River) was recently remediated by dredging 2.7 million cubic yards of sediments, processing, and shipping via train to approved landfills at a huge cost. Prediction of microbial bioremediation of PCB-contaminated sediments has been challenging due to a lack of understanding of the true kinetic limitations and the difficulty in measuring the microbial kinetics at environmentally relevant aqueous concentrations. This contrasts with the parallel developments in microbial dechlorination of chloroethenes (CEs) in groundwater, which share similar microorganisms and mechanisms. Many groundwater sites contaminated with CEs have been remediated using in situ bioaugmentation for dechlorination. We propose that key differences for PCBs compared to CEs are lower aqueous concentrations (lower rates of dechlorination), stronger hydrophobicity and binding to solids (lower bioavailability), need for both anaerobic and aerobic processes (to achieve mineralization), and limited knowledge of the rate-limiting steps (mass transfer vs degradation).

Figure presents a historical perspective of work on the microbial degradation of PCBs. Following a few early findings of PCB microbial degradation in the 1970s, evidence emerged in the 1980s of highly diverse microbial pathways that dechlorinated PCB mixtures to less chlorinated products that were subsequently aerobically degraded in the environment. , Much of this evidence was observed as shifts in chromatography peak patterns. The pace of these studies increased in the 1990s via analysis of multiple sites, different research groups, and identification of pathways. , In the first decade of the 2000s, more pathways were identified, a microbial strainwas linked to PCB dechlorination, and a PCB-respiring bacterium was isolated and grown in the laboratory. While the first evidence of microbial dechlorination came from investigating PCB contamination in the field, the work in the subsequent three decades remained in the laboratory and paths to field-scale bioremediation remained elusive. Critical bottlenecks for scaling had not yet been identified. Processes that potentially limited native rates of dechlorination included availability of dechlorinators, electron donors, nutrients, and most importantly, the bioavailability of PCBs in sediments. What was clear to researchers was that microbial dechlorination occurred at nearly every site evaluated, but the rates were too slow to make a material difference within preferred time scales of years rather than decades or centuries. Confirmation that microbial dechlorination kinetics were governed by dissolved PCB concentrations provided a physiological explanation for low numbers of PCB-dechlorinating bacteria in sediments and led to the possibility of addressing the challenge through bioaugmentation. The subsequent work in the second decade of this century involved engineering scale-up of dechlorinator growth in large reactors, methods for large-scale delivery into sediments, and demonstrations of PCB bioremediation at multiple field sites. These efforts were aided by some synergies with the parallel development and testing of in situ remediation of PCB-impacted sediments with activated carbon.

1.

Historical perspective of key developments and published papers in the field of PCB microbial dechlorination and aerobic degradation. Using a PubMed advanced search, 369 articles were found for PCB dechlorination, 184 for aerobic PCB degradation, and 23 for both. The search was performed using the terms: PCB or polychlorinated and dechlorination; and PCB or polychlorinated and aerobic. Articles not involving microbial dechlorination/degradation or focusing on other chlorinated compounds were removed.

In the discussion below, we present a historical perspective of key developments in the field generally following the timeline of critical milestones illustrated in Figure . We conclude with a perspective on the path forward toward successful scaling of PCB bioremediation to achieve the goal of a sustainable and low-cost approach for remediation of PCB-contaminated sediments.

Early Findings from the Field

Microbial dechlorination of PCBs was first observed in highly contaminated sediments in the upper Hudson River downstream from a General Electric facility in New York. In the summer of 1991, six cylindrical and sealed caissons (5 × 1.8 m²) were lowered into the water column and sediment of the Hudson River, 312 km upstream of New York City. Thus, began the first attempt to bioremediate PCBs in situ. The primary approach was to utilize indigenous aerobic bacteria to degrade the mono- and dichlorobiphenyl congeners that had accumulated due to what at the time was suspected to be the in situ anaerobic dechlorination of Aroclor 1242. Several laboratory studies preceded the aerobic bioremediation field trial demonstrating that lesser chlorinated congeners could be converted to chlorobenzoates and then mineralized by aerobic bacteria. , Similar aerobic microbial degradation of the PCBs in situ was observed during the field trial in 1991 with further enhancement of the biodegradation achieved following amendments of oxygen, biphenyl, and nutrients. Overall, the aerobic field trial successfully removed lesser chlorinated PCB congeners from the environment. As illustrated in Figure , research on aerobic and anaerobic processes continued in parallel throughout the last five decades. However, more extensively chlorinated PCB congeners, generally with five or more of the 10 possible chlorines, were found to be resistant to aerobic biodegradation, which hindered the bioremediation of commercial mixtures of PCBs, but such congeners could be dechlorinated under anaerobic conditions. It was clear at that time that to develop a robust in situ treatment for PCBs in sediments, both anaerobic and aerobic microbial transformations would be needed. Anaerobic dechlorination was still poorly understood, and how to combine the redox conditions, render the PCBs available to the microbes, and scale and implement such a process was still far in the future. The next step was to understand the microbial PCB dechlorination.

Discovery of Microbial Dechlorination Pathways

As early as 1974, dechlorinated metabolites were observed in rabbits fed 2,2′,4,4′,5,5′-hexachlorobiphenyl. In 1987, evidence indicating that highly chlorinated PCBs could be dechlorinated under anaerobic conditions in situ was reported for multiple Aroclor mixtures of PCBs in sediment at several sites. A surge of laboratory studies through the 2000s revealed that microbial PCB dechlorination was not an isolated phenomenon, but rather was widespread with diverse outcomes in anoxic environments spanning sediments in ponds, rivers (in addition to the Hudson), − lakes, , marine and estuarine systems, , and waste treatment systems. ,

Discovery of PCB dechlorination was contemporary with discovery of dechlorination of the CEs, another widely distributed class of contaminants. , However, harnessing PCB dechlorination would be a monumental challenge compared to CEs because not only were there dozens of possible PCB dechlorination products to contend with, compared to just a few resulting from chloroethene dechlorination, but PCBs were more hydrophobic, nearly insoluble, and co-planar PCB congeners impart “dioxin-like” toxicity with additional regulatory implications for remediation.

Given the complex soup of possible daughter products, bringing understanding to PCB dechlorination processes in diverse environments needed a systematic approach. The challenge was “how” when neither the organisms nor the responsible enzymes were known. When Brown et al. examined chromatograms from Hudson River sediments, they noted stunning complexity with dozens of congeners indicating, as they determined, diverse dechlorination daughter products being produced from the original contaminating Aroclors. The scientists astutely distilled and classified the chromatographic observations into recognizable and recurring patterns of dechlorination products produced from original mother congeners. They introduced letter-designated pathways, then described in exquisite detail additional processes as new environments were studied, , and PCB-dechlorinating strains were identified. Ultimately, nine separate pathways from process “H” to “Z” were identified.

In the absence of isolated strains, pattern ID enabled hypotheses that specific bacterial species or separate enzymes were responsible for activity leading to the production of different suites of congeners. It was not until 2007 that a heroic effort by Fagervold et al. conclusively confirmed through quantifying 16S rRNA genes that specific PCB-respiring phylotypes in a mixed community mediated dechlorination of specific PCB congeners via specific preferred chlorine position removal. This groundbreaking study also showed growth on PCB congeners and brought PCB dechlorination to a similar knowledge base as CE dechlorination. As Bedard et al.’s process letters indicated, the organisms and enzymes likely varied according to the environment where PCB pollution had originally occurred and exerted different dechlorination patterns.

How PCB dechlorinators existed in complex microbial food webs was unclear for some time since the presence of competing electron acceptors, most notably sulfate, found in abundance in estuarine and marine environments, had different effects. It either inhibited PCB dechlorination, ,, showed no inhibition, or led to stimulation of dechlorination after sulfate was depleted. Indeed, Zwiernik et al. theorized that sulfate-reducing bacteria were responsible for PCB dechlorination in that study and that, after depleting the initial sulfate electron acceptor, they switched to PCB electron acceptors.

Isolation of the first Dehalococcoides on PCE (strain 195) and later, on trichlorobenzene (strain CBDB1), along with identification of multiple reductive dehalogenases in Dehalococcoides genomes advanced organohalide respiration as an obligate mode of microbial existence, and as a potential bioremediation tool for organohalide pollutants. Examination of Dehalococcoides strains for specific activity on PCBs soon followed. , A lab-based PCB bioaugmentation study with Dehalococcoides mccartyi strain 195 , showed increased PCB dechlorination, but with decreasing abundance of the strain over time, in sediments. Molecular detection of 16S rRNA genes and reductive dehalogenases (Rdhs) was effective for monitoring those organisms and functional genes in the active treatments. Dehalococcoides strain JNA isolated from PCB-contaminated sediments by Bedard’s group was confirmed to grow on PCBs.

Isolation of Dehalococcoides (strains CG1, CG4, and CG5) yielded, for the first time, identities of Dehalococcoides genes encoding for Rdhs linked to PCB dechlorination. Bedard, an original author on the earliest papers about PCB dechlorination, connected the discovery of PCB Rdhs to dechlorination patterns (H to Z) – i.e. multiple Rdhs specific for specific PCBs provide a mechanistic basis for those patterns that were empirically defined decades earlier. Advancement continues with a combination of machine learning, enzyme assays, and quantum chemical calculations being recently applied to predict the outcomes of dechlorination of all PCB congeners.

Although Dehalococcoides are now widely used as bioaugmentation agents in groundwater CEs, isolation of Dehalobium chlorocoerica DF1, a PCB-dechlorinating member of a closely related genus, fueled advancement of bioaugmentation for PCBs bioremediation.

Advancement with New Microbiology and Molecular Biology Tools

An early challenge to identifying PCB-dechlorinating bacteria was that they could not be isolated by colonization on a solidified medium. An additional difficulty was that they only grew in sediment microcosms, making it difficult to separate them from faster-growing community members using serial transfers. PCB-dechlorinating bacteria were first identified in 2001 by combining 16S rRNA gene amplification using universal bacterial primers during PCB dechlorination. This was followed by isolation of a novel organism, Dehalobium chlorocoerica DF-1, in coculture with a Desulfovibrio sp.

To advance the use of dechlorinating bacteria for PCB bioremediation, the development of methods for detecting them was critical. Although 16S rRNA gene probes employed in polymerase chain reaction (PCR) are instrumental for characterizing environmental bacteria, they may not distinguish between dechlorinating strains because their 16S rRNA gene sequences are highly conserved, and they do not specify growth on different organohalides, e.g., PCBs. Furthermore, 16S rRNA genes are not ideal for monitoring PCB-dechlorinating bacteria added for bioaugmentation in PCB-impacted sites because they may not differentiate between indigenous and added strains. Thus, detection and enumeration of functional Rdh genes that encode PCB-dechlorinating enzymes were found to be more useful.

Payne et al. detected a putative reductive Rdh gene from DF-1 and used DNA primers of the conserved flanking regions of the gene sequence for quantitative PCR to monitor DF-1 in Aroclor 1260-contaminated sediment mesocosms. Burkholderia xenovorans LB400, added concurrently with DF-1 to aerobically degrade less chlorinated PCB congeners, was monitored using a conserved DNA sequence upstream of the bph operon encoding biphenyl dioxygenase for quantitative PCR. ,, Using this approach, both strains (DF-1 and LB400) were selectively quantified by quantitative PCR in mesocosms and in the field. , Now, with the identification of the genes for three PCB Rdhs, pcbA1, pcbA4, and pcbA5, each unique to three strains of D. mccartyi and with distinct specificities, it might be possible to use all three strains for bioremediation and monitor each strain to determine if one is dominant or if there is a synergistic effect among the strains. Moreover, genome-resolved metagenomic and metatranscriptomic analyses of a contaminated site revealed additional Rdhs, including a previously unknown expressed putative PCB Rdh associated with Dehalococcoides, suggesting that the diversity of PCB-dechlorinating enzymes and potential is even greater than currently recognized.

The current ability to selectively monitor bioaugmented PCB-dechlorinating bacteria and differentiate them from indigenous PCB dechlorinators provides an essential tool for assessing the effectiveness of bioremediation at different PCB-impacted sites.

Measuring Rates of Dechlorination and Identifying Bottlenecks

Despite the widespread presence of PCB-dechlorinating bacteria, dechlorination of PCBs in the natural environment was slow, and impediments were not well understood. For example, in a natural attenuation study for Lake Hartwell, SC sediments, Pakdeesusuk et al. observed rates in 1998 that were slower than rates predicted in 1987, suggesting a plateau in dechlorination activity. Similar plateau effects were reported for the St. Lawrence River, Grasse River, and Hudson River sediments. − In each of these studies, the dechlorination rate was based on sediment PCB concentration, and the PCB concentration in porewater was not measured. Slow dechlorination in the field appeared to be due to limited bioavailability of PCBs to the bacteria. Zwiernik et al. proposed that residual petroleum in sediments reduced the rate of PCB dechlorination by reducing the bioavailable fraction in the freely dissolved phase; however, their rate studies did not include dissolved PCB concentrations. Two types of kinetic thresholds for PCB dechlorination were also reported. The first type is that even under optimal conditions for dechlorination, the rates slow when the residual concentration falls below a threshold value. Kinetic studies of Lake Hartwell sediments , and St. Lawrence River sediments observed this plateau effect but could not provide an explanation for the observations, attributing this effect to a kinetic threshold for PCB dechlorination. The second threshold effect pertains to sediment PCB concentrations that are unable to support the growth of the dechlorinating bacteria to reach cell numbers that can substantially reduce sediment PCB concentrations within a reasonable time period. Sediment slurry studies suggested growth-related kinetic thresholds for PCB dechlorination as high as 40 mg/kg.

Lombard et al. measured PCB dechlorination kinetics as a function of freely dissolved concentration using a novel passive dosing-passive sampling approach and found no concentration threshold for anaerobic dechlorination of PCB 61 by DF-1 at environmentally relevant aqueous concentrations (1–500 ng/L). PCB dechlorination exhibited a first-order aqueous rate constant, regardless of the initial aqueous concentration. Needham et al. demonstrated that the rate of dechlorination was linearly dependent on the freely dissolved concentration of PCBs both in sediment and in sediment-free microcosms, and the rate of dechlorination in sediment microcosms could be predicted within a factor of two based on the kinetics measured in sediment-free microcosms. They suggested that the low PCB dechlorination rates observed in weathered sediments are due to either low native cell densities or slow desorption kinetics. They showed that at a native cell abundance of 10³ cells/g it would take decades to significantly reduce PCB concentrations in field sediments as generally observed (Figure ). Only when the cell abundance was increased above 10⁵ cells/g was there a reduction of the PCB half-life to years instead of decades, and the rate was limited by desorption and not dechlorination. Thus, a major breakthrough in PCB bioremediation is the possibility to overcome the bottleneck of growth threshold for PCB dechlorination through engineered bioaugmentation to increase cell numbers, as has now been demonstrated in laboratory , and pilot studies. It is important to note, however, that modeling of kinetic limitation has not been performed in field sites.

2.

Effect of cell density of Dehalobium chlorocoercia DF-1 on dechlorination rate compared with the rate of desorption of PCB 194. The yellow shaded region depicts the range between fast and slow desorption rates from sediment estimated for Grasse River sediment.

However, even in bioaugmented systems, there is an optimal level of bioaugmentation, and the final extent of dechlorination will be controlled by the relative kinetics of microbial dechlorination and slow desorption from sediments, and the proportion of the slowly desorbing PCB pool in sediments. In organic-rich anaerobic sediment environments where PCBs are often found, there should be sufficient availability of electron donors that are not limiting. Once the pool of fast-desorbing PCBs in sediment has been exhausted by a microbial population, it is possible that the population will decrease as less PCBs are available in the aqueous phase and are replenished by the slow-desorbing pool. This population “die-off” in bioamended sediments has been observed in both microcosm and pilot studies in which the cell density decreases by 1 to 3 orders of magnitude after one year. ,

Mass Culturing and Synergy between Aerobic/Anaerobic Processes

A significant barrier to bioaugmentation for bioremediation of PCBs is scaling up the growth of sufficient biomass for treatment of contaminated sediments. PCB-dechlorinating species grow slowly with PCBs and do not grow to high culture densities. Fortuitously, all described PCB-respiring species also grow with CEs such as perchloroethene and trichloroethene. CEs offer several advantages for growing PCB-dechlorinating microorganisms: (1) CEs support growth that is 5 times faster and results in 10–20 times greater cell densities compared with PCBs, (2) CEs are highly volatile and unlike PCBs, they can be purged from cultures by N₂ sparging prior to harvesting, (3) there is negligible lag in growth when CE-grown cells are transferred to PCBs. Commercial dechlorinating cultures (e.g., SDC-9 and KB-1) are routinely grown as mixed consortia for bioremediation of CEs in groundwater. Dehalobium chlorocoercia DF-1 has been successfully mass cultured for bioremediation of PCB-contaminated sediments using a similar approach. In contrast to Dehalococcoides mccartyi strains 195, CBDB1, CG1, CG4, and CG5 that require hydrogen as an electron donor, DF-1 also utilizes sodium formate. ,, Sodium formate is highly soluble and noncombustible, which is advantageous for high-volume bioreactors where mass transfer of hydrogen gas can be limiting. The maximum biomass yields of DF1 are similar to those obtained for pure and mixed cultures of D. mccartyi strains in the range of 1–2 × 10⁸ cells mL^–1. At these yields, a 200 L bioreactor generates sufficient biomass to treat 4000 m² of sediment. Although scale-up costs for treating larger areas are lower than other treatments such as dredging, identifying factor(s) that limit culture density could further reduce the treatment cost for large, contaminated sites. Since culture viability decreases after storage for 1–3 months, another limitation to be considered is the long-term preservation of cell biomass. Amendments to enhance growth and activity of dechlorinating bacteria have long been sought and include cobalamin, a required co-factor for Dehalococcoides; surfactants, e.g., rhamnolipids, which have enhanced rates of dechlorination of PCBs in sediment microcosms; and resuscitation-promoting factor from Micrococcus luteus, which enhanced rates of dechlorination of PCE and PCBs. These approaches could advance not only cultivation and storage, but also sediment deployment of PCB-dechlorinating bacteria as bioaugmentation agents.

Bioremediation of PCBs requires complementary activities of both anaerobic dechlorinators and aerobic biphenyl-degrading bacteria for complete mineralization. Sequential anaerobic–aerobic treatment has been reported in laboratory studies with soil and sediment microcosms containing both weathered and spiked PCBs. − Master et al. successfully demonstrated degradation of weathered Aroclor 1260 in soil microcosm by sequential anaerobic dechlorination by indigenous dechlorination and aerobic degradation with B. xenovorans LB400. Long et al. reported that this approach was potentially feasible for ex situ treatment of composted contaminated soil in a composting reactor. However, sequential anaerobic–aerobic treatment was thought not to be feasible for in situ bioremediation of highly anoxic sediments.

A tractable approach for treating PCB-contaminated sediments was first reported by Payne et al. Open mesocosms containing sediment contaminated with Aroclor 1260 were bioaugmented with both cultures of D. chlorocoercia DF-1 and the aerobic biphenyl degrader B. xenovorans LB400, and this resulted in >80% reduction in the total mass of PCBs after 365 days. The results demonstrated that anaerobic and aerobic processes occurred concurrently in an open system, likely as a result of benthic activity that caused dynamic shifts in redox potential throughout the benthic zone of the sediments. Biofilms form redox gradients on sediment particles that protect aerobes and anaerobes during fluctuations in the redox potential. While laboratory feasibility was demonstrated, methods for large-scale field deployments were yet to be developed.

Synergy with In Situ Remediation with Activated Carbon and Field Demonstrations

The advancement of in situ remediation of PCB-contaminated sediments with sorbent amendment and development of methods for large-scale broadcasting of treatment amendments into sediments , provided opportunities for synergistic applications combined with bioaugmentation for PCB degradation. Activated carbon is applied to the sediment matrix to reduce porewater PCBs and prevent PCBs from bioaccumulating and biomagnifying in the food web. This technology provides a paradox for AC impact on microbial dechlorination: on one hand, the addition of AC can impede microbial dechlorination kinetics, as it reduces porewater PCB concentrations, , which could result in slower PCB dechlorination kinetics in the long term. Alternatively, AC can enrich PCBs on surfaces which can be colonized by PCB degraders enhancing proximity and potential dechlorination, as illustrated in Figure . However, proof of the relative impacts of these mechanisms on PCB microbial degradation is scarce. The rough surface of AC provides a microhabitat for PCB-degrading microorganisms, whether aerobic/anaerobic, for their requirement of a specific redox environment. Both aerobic PCB degraders and anaerobic dechlorinators can be inoculated onto the AC, − and both aerobic degradation and anaerobic respiration can be carried out. ,, AC has also been observed to promote the activities of quorum-sensing genes to upregulate biofilm potential and cell–cell communication pathways. Numerous studies have reported that AC can serve as an electron-mediating matrix to promote electroactive dechlorinating bacteria to dechlorinate various organohalides, − but related studies on PCBs are scarce and inconclusive. Several published papers have suggested potential synergistic effects. First, AC can serve as the delivery vehicle for microbes to reach the sediment, which itself is a substantial engineering challenge. Second, the dechlorinators seem to be able to access and dechlorinate PCBs in the presence of AC. Recent work by Dong et al. further stabilized PCB-degrading LB400 biofilms on biochar and continued to observe degradation activity. Thus, both aerobic and anaerobic activities seem to proceed on biofilms attached to strong sorbents.

3.

PCB-dechlorinating biofilms on activated carbon enable convenient delivery of microorganisms for bioaugmentation, enabling long-term bioremediation along with immediate reduction of PCB exposure to ecosystems.

Several pivotal studies published in ES&T have successfully scaled this new technology from batch to field (Figure ). ,, A small pilot-scale demonstration was conducted at a watershed drainage creek containing an average of 5 ppm Aroclor 1254/1260 to assess the efficacy of using natural bioturbation rather than active mixing of the bioamendment (activated carbon inoculated with PCB-dechlorinating and degrading organisms) into the sediment. Treatments with the bioamendment in 400 m² plots achieved up to 52% reduction in total PCB levels in sediments and 95% reduction in porewater PCB concentrations in the bioactive zone of 15 cm 13.5 months after treatment. No significant changes in sediment PCBs were observed in untreated and AC-only sediment plots. The results demonstrated that active mixing of the bioamendment into the sediments was not required for the effective distribution of the bioamendment. The study confirmed the feasibility of using bioamended AC for the in situ treatment of PCBs in contaminated sediments. Several other successful pilot-scale applications have been performed at additional sites, including an ongoing NIEHS SBIR Phase 2 demonstration at the New Bedford Harbor Superfund site. As results from these studies in the field emerge and positive results are demonstrated, in situ bioaugmented sorbent amendment may be one of the few economically feasible options in PCB sediment remediation: rapid binding to AC and exposure reduction, and long-term microbial dechlorination and degradation. Techno-economic evaluations will have to be performed on a site-specific basis to evaluate the appropriateness for each site.

Our knowledge of how microbial dechlorination and AC synergy work is still limited, thus restricting model development and the prediction of effectiveness. Related literature on modeling tools coupling sorption and biotransformation of PCBs on AC is scarce. Furthermore, how PCB degraders colonize the AC surface is not well understood, which poses challenges in comparing different experimental results of the batch system. These are the factors that need to be considered to develop an accurate engineering model for the progress of bioaugmented sorbent amendments.

Current State of the Science, Progress with Remediation, Steps Forward

As a collective achievement of the PCB research over the last 50 years, research on PCB bioremediation has moved from the laboratory to the field. Much progress has been made in understanding the underlying biological processes and limitations of the physical environments of PCBs in sediments. Several pilot-scale studies have now demonstrated the feasibility of inoculating AC with aerobic and dechlorinating cultures together before delivery into sites by using large-scale equipment. The delivered organisms survive the journey, as detected by molecular monitoring, and significantly enhance the rate of transformation of PCBs in the field. Importantly, in conjunction with the AC amendment, ecosystem exposure to PCBs is rapidly reduced, while enhanced PCB degradation continues over the long term.

Work is needed for scaling up the production of dechlorinators along with the short- and long-term preservation of cells to provide on-demand availability for large-scale projects. Maximum growth densities of dechlorinators in reactors are still limited, and further increasing cell densities will reduce the cost. Other approaches of combining aerobic and anaerobic degraders may need to be tested in parallel, including some of the innovative approaches of stabilizing/protecting biofilms with advanced material engineering.

PCB dechlorination is an inherently slow process, and long-term field studies are needed to completely understand limitations of field microbial kinetics vs desorption from sediments. Additionally, the need for multiple timepoint bioaugmentation needs to be evaluated in the field as cell numbers are observed to decline over time.

While the focus of the larger scientific and engineering community has shifted to address new and emerging contaminants in the environment, we still need to rid our waters of damaging PCBs that continue to impact our natural resources and human health. Many individuals and entities have encouraged and funded decades of critical research that has now led to the first field applications of PCB bioremediation. While the progress has been significant, support for continued development of economical and sustainable PCB bioremediation technologies is needed to achieve large-scale clean-up of the many remaining PCB-contaminated areas. Decades of research to identify the factors that limit degradation by natural attenuation in the environment led to a tractable approach for bioremediation of PCBs, once thought to be highly recalcitrant in the environment. This approach serves as a model for remediating other legacy and emerging “forever chemicals” that have accumulated in the environment.

Biography

Dr. Upal Ghosh is a professor in the Department of Chemical, Biochemical, and Environmental Engineering at the University of Maryland Baltimore County. He has an undergraduate degree in Chemical Engineering and MS and Ph.D. degrees in Environmental Engineering. His group performs research in environmental engineering and science with a focus on the fate, effects, and remediation of toxic pollutants in the environment. His fundamental research has contributed to the development and transition of nature-based sediment remediation technologies that work by altering sediment geochemistry and enhancing biological degradation. Dr. Ghosh has also led the development of advanced monitoring tools for pollutant bioavailability, especially work on passive sampling of pollutants in water. Technologies developed by his group are being implemented at full scale to provide sustainable solutions for large-scale coastal restoration cum remediation efforts. His work has been published in leading journals, US patents, and recognized through multiple awards. Dr. Ghosh is also the co-founder of two startup companies that are transitioning emerging sediment remediation technologies to the field. He works closely with industry and regulatory leaders to bring the best science into the practice of Environmental Engineering.

The authors declare the following competing financial interest(s): Several authors are co-inventors of patents related to the technologies described in this paper for which they are entitled to receive royalties. KS and HM are co-inventors of U.S. Patent No. 6,946,248 and 7,462,480 B2 issued to the University of Maryland Baltimore County (UMBC) and Medical University of So. Carolina. KS ad UG are co-inventors of U.S. Patent No. 8,945,906 issued to UMBC. UG and KS are partners in a startup company (RemBac Environmental) that has licensed the three technologies. UG is also a co-inventor of US Patent No. 7,101,115 B2 issued to Stanford University and U.S. Patent No. 7,824,129 issued to UMBC. UG is a partner in a startup company (Sediment Solutions) that has licensed these two technologies from Stanford and UMBC and is transitioning to the field.

Published as part of Environmental Science & Technology special issue “60th Anniversary of Environmental Science and Technology”.