For petroleum fuels and crude oil, natural biodegradation processes and rates have been investigated for decades [70, 71], and for plumes [29, 72] outcomes from groundwater investigations are now formalized in national guidelines in the USA, Australia, and the United Kingdom– often termed Monitored Natural Attenuation (MNA) [73, 74, 75]. MNA as depicted in Fig. 2 is inclusive of all attenuation mechanisms including dispersion (leading to lower concentrations), sorption (usually leading to delays in the transport of contaminant mass in groundwater), and biodegradation due to co-consumption of electron acceptors (leading to true mass loss or conversion within the groundwater). Challenges to implementation of MNA for petroleum hydrocarbons are (i) the need to demonstrate stable or shrinking plumes that are not crossing jurisdictional boundaries; (ii) that risks to human health and the environment are acceptable, e.g., from compounds such as benzene, and (iii) that data and information is of such a quality to enable quantitative MNA assessment.

Often MNA approval for management of a petroleum hydrocarbon plume also requires removal of the source of the plume, which is usually a non-aqueous phase liquid (NAPL) product held in the vadose zone or within the near-depth interval of the zone of water table fluctuation. Natural Source Zone Depletion (NSZD) is now often being used to manage LNAPL in the subsurface where NSZD rates are deemed high enough compared to active source removal techniques [76]. NSZD is revisited later in the paper when considering biodegradation in the context of holistic approaches to plume management.

Chlorinated organic compounds such as PCE, TCE, 1,2 dichloroethane (1,2 DCA) and others can be degraded to different degrees both aerobically and anaerobically. Outcomes of many early biodegradation studies are summarized in Wiedemeier et al. (1999) [29]. A strong focus has been reductive de-chlorination and the role of electron donors (such as natural organic carbon or carbon amendments such as toluene, acetate, hydrogen, emulsified vegetable oil [77, 78, 79]) in driving de-chlorination to ethene/ethane [80]. More specialized amendments such as poly-3-hydroxybutyrate have been evaluated to assess persistent release and supply of electron donors [81]. The organisms performing the de-chlorination are using the chlorinated compounds as electron acceptors to sustain growth [82]. Largely, the only microbes that have been found to fully degrade chlorinated ethene organics to ethene are members of the group Dehalococcoides [83].

A concern has been the sometimes lack of complete mineralization leading to the accumulation of the potentially more hazardous compound vinyl chloride (VC) as a degradation product. Studies have also shown that the less chlorinated hydrocarbons such as 1,2 DCA and VC do degrade under aerobic conditions both in laboratory and field studies [51, 84, 85]. Formalized national guidance in the United States integrates the information and approaches needed to assess the natural attenuation of chlorinated solvent groundwater plumes for regulatory compliance [86], but similar guidance does not seem available elsewhere. Whilst we have come to understand the drivers for degradation of chlorinated hydrocarbons in groundwater, the variability of redox conditions and electron acceptors vs electron donors needed to maintain biodegradation processes compared to petroleum compounds, provides greater uncertainty for regulators to sign off on MNA of chlorinated hydrocarbon impacted sites.

Pesticides and their precursors such as atrazine and chlorophenols [6], like chlorinated solvent compounds, are xenobiotic, being man made and largely introduced to the biosphere, compared say to petroleum compounds which are naturally sourced and somewhat ubiquitous. As such, microorganisms that come into contact with xenobiotic compounds may not initially have the capability to degrade them. For atrazine, such observations were made by Franzmann et al. (2000) [87] who identified that atrazine degrading microorganisms needed to be isolated and augmented into atrazine contaminated aquifer material to induce biodegradation of the atrazine. Under aerobic conditions atrazine was shown to be fully mineralized [50].

Other organic compounds commonly contaminating groundwater and found to biodegrade and transform are munition compounds, and for chemicals used in Defense applications such as perchlorate. For munition compounds, 2,4-dinitrotoluene (2,4-DNT) altered microbial communities during biodegradation [88], the potential for biodegradation of RDX (hexahydro-1,3,5-trinitro-1,3,5-triazine) was evaluated using a combination of differing redox incubation conditions [48]; and degradation of trinitrotoluene, 2,4-DNT, and 3-nitrotoluene was shown to occur in microcosms, with biodegradation occurring under nitrate-reducing conditions whilst creating intermediate metabolites under aerobic conditions [47]. For perchlorate (typically used in rocket fuel), prior research on chlorinated solvents enabled rapid advances in its understanding and treatment by targeting the use of electron donors to accelerate biodegradation rates [89]. An overview is given in Stroo and Ward (2009) [90].

A range of other emerging chemicals are also being actively investigated to determine their biodegradability in groundwater [91]. As mentioned, research indicates that some of the organic compounds that comprise the precursors to PFAAs will bio-transform though a recent review states “they cannot be degraded completely through a single biodegradation pathway” [92]. Also, biotransformation within PFAS plumes is a process that may not expedite remediation but may exacerbate remediation prospects by converting often lower mobility unregulated PFAS precursors to higher mobility, regulated PFAAs [93].

The nitrogen cycle is well known, and there is a long history of research on nutrient biotransformation in groundwater especially that of nitrogen species such as nitrate and ammonium [94]. Significant contamination of groundwater has occurred due to agricultural practices, landfills, urban sewage, industrial use and other sources [95]. Denitrification of nitrate (transformation of nitrate to nitrogen gas) commonly occurs where available carbon sources are present [96], although ammonification (conversion to ammonium) may sometimes occur. Denitrification may also occur in the presence of other reduced species such as sulfides [97]. Ammonium, in contrast, is stable under reduced geochemical conditions but under aerobic conditions will transform into nitrite and nitrate [96].

As with any contaminants (and metals in particular) general attenuation occurs due to dispersion but also by sorption to aquifer sediment surfaces or potential binding with mobile organic carbon. A range of metals can change their chemical status and behavior based on aquifer geochemical conditions [98]. Microbial reduction of metals such as chromium and uranium has been shown to produce both reduced soluble, mobile end products as well as insoluble non-motile end products [99, 100, 101]. Attenuation processes for a number of inorganic contaminants such as copper, lead, cadmium, and chromium in groundwater have been summarized [102]. Bioprocesses can alter pH and other geochemical conditions to induce sorption/desorption processes that might bind or release some metals, and bacterial sulfate-reduction can also lead to bioprecipitation of resultant insoluble metal sulfides onto aquifer sediments removing metals from groundwater plumes [103, 104].

Because of limited mixing and the laminar flow nature of groundwater systems liquid injection of amendments (be they electron acceptors/donors, nutrients, acidity, etc.) into aquifers to induce greater biodegradation rates may result in an isolated volume of the amendment in the aquifer due to hydraulic displacement of the targeted groundwater plume away from the well location(s) during the fluid injection phase. To overcome this and to enhance biodegradation and remediation, several strategies have been considered including (i) pulsed injection of liquid amendments to induce greater surface area contact zones between the injected volumes and the contaminant plume [105, 106]; (ii) recirculating groundwater pumping wells whereby wells (or screens within the one well) might extract and inject groundwater with added amendments prior to groundwater reinjection or to induce mixing [107, 108, 109]; (iii) establishment of a stationary or less mobile treatment zone in an aquifer across the groundwater plume flow direction that reacts with the contaminant of interest – such as permeable reactive barriers (PRBs) [79, 110, 111]; (iv) injection of reactive gases (such as oxygen, hydrogen) that might accelerate biodegradation but also might remove mass by other means [51, 112]; (v) bio-electrochemical systems which can stimulate the interaction of microbial metabolism with electrodes and can aid physical migration due to the electric field generated [113] and (vi) injection of slurry or placement of solid or particulate material to induce greater biodegradation [114, 115]. Advantages and disadvantages of some of these methods are collated in Table 1 (Ref. [106, 116]). In short, all methods are challenged by the often large scale of pollutant plumes in groundwater, because if plumes are greater than 100s of metres long and wide, close spacing of injection/extraction infrastructure or ongoing pulsing of injections might be prohibitively expensive. Rebound and back-diffusion effects can also prolong groundwater remediation timeframes. PRBs and other ‘across-plume’ injection or amendment placement strategies often enjoin fewer initial costs but only treat groundwater at selected cross-plume locations.

Despite the limitations mentioned in Table 1 across groundwater bioremediation strategies, below we provide a brief overview of two primary strategies and selected field case studies for PRBs and injection of gases, that highlight their potential, inclusive of some outcomes for bioaugmentation.

Bioaugmentation is an additional strategy for enhancing biodegradation [83]. Bioaugmentation is the addition of microbes into an aquifer to effect remediation and could be implemented as part of a number of the strategies in Table 1. The aim of bioaugmentation is to introduce specialized metabolic capabilities to induce biodegradation or transformation of specific compounds, or to bolster population sizes where they are limited and by adding other amendments alone may not increase the relative abundance of the desired biota. To be effective the augmented microbes must survive, grow, and retain their degradative capacity. Note that in many cases addition of microbes may not be needed since indigenous biota in groundwater are often capable of degrading contaminants if they have been resident for some time; geochemical and mass transport conditions usually limit biodegradation processes. One of the advantages of bioaugmentation is the ability to select the desired metabolic activity. This can be particularly beneficial where microbes might take considerable time to gain the capability to mineralize contaminants of interest (e.g., atrazine [87], and see the latter discussion). And, while bioaugmentation has become a more common treatment for sites contaminated with chlorinated solvents, it has not performed as well for other pollutants [83]. Bioaugmentation may be more successful in ex-situ systems such as bioreactors treating contaminated soil or groundwater. Disadvantages of bioaugmentation are that few successful field demonstrations have occurred beyond chlorinated solvents, it has been difficult (but becoming more feasible [117]) to monitor and differentiate introduced species from indigenous microbes, introduced species may be outcompeted by natural microbes upon addition to subsurface systems (in particular), microbes may lose their degradation capability upon injection, and transport and delivery of microbes to the target regions of an aquifer may be problematic. Regardless, via rapid progress in molecular biology capabilities and the genetic manipulation of microorganisms, new gene bioaugmentation holds promise [118].

Davis and Patterson (2003) [119] reviewed progress and developments with bio-treatment PRBs. They outlined operational needs and the potential for amendments into PRBs that could biodegrade and/or bio-transform petroleum hydrocarbon, ammonium/nitrate, chlorinated hydrocarbons, metals and other contaminants in groundwater. A number of studies have investigated the use of slow-release carbon sources in PRBs (e.g., sawdust [120, 121], cotton [122], peat [123], imitation vanilla [124], vegetable oil [79], hydrogen/methane [125], leaf mulch/wood chips [126], mulch [127]) for reductive bioremediation of contaminants such as nitrate, chlorinated solvents, munitions (e.g., RDX), sulfate (and low pH) and some metals in groundwater. Some PRBs have used oxygen release compounds placed across the pathway of BTEX plumes. Bianchi-Mosquera et al. (1994) [128] evaluated peroxide briquets to provide an electron acceptor for the biodegradation of injected benzene and toluene, Borden et al. (1997) [129] used immobilized nutrients and oxygen in briquet form; and Kao and Borden (1997) [130] used slow-release nitrate but observed that of the BTEX compounds benzene did not biodegrade. Apart from being able to identify the correct amendment that would create suitable conditions for enhanced biodegradation, linking groundwater flow through PRBs and ensuring adequate residence time for reaction kinetics is important for effective performance. Ensuring reduced hydraulic conductivities are not induced due to amendment emplacement within the PRB or through PRB clogging is also critical to not divert groundwater flows around the PRB treatment zone. To provide greater control on delivery rates and the spread of amendments in PRBs hollow fiber membranes and polymer tubes woven into mats have been evaluated in laboratory and field conditions [50, 96, 131].

The PRB scenarios depicted in Fig. 3 (Ref. [50, 96]) are for an atrazine plume [50] and an ammonium plume [96, 132]. For the atrazine PRB example (Fig. 3A), atrazine degrading microorganisms were isolated and bioaugmented into groundwater where atrazine biodegradation was not occurring [87], and to sustain the microbial community and stimulate bioremediation of the atrazine plume, oxygen was supplied by passing air through a network of polymer mat tubing providing an augmented and near constant oxygen supply [50]. A primary result is shown in Fig. 3C, indicating rapid biodegradation when oxygen was supplied via air injection through tubes woven into a polymer mat placed across the flow direction. The strategy was shown to be successful in achieving a short degradation rate half-life of about 0.35 days, compared to no biodegradation without oxygen amendment.

Fig. 3.

(A) Schematic of atrazine biaugmentation treatment barrier. In this scenario, an atrazine plume is being remediated by injection of atrazine-degrading bacteria and delivery of oxygen creating a permeable biobarrier wall at the leading edge of the plume. (B) Schematic of a dual (sequential) PRB to treat an ammonium plume - (i) to deliver oxygen to form nitrite/nitrate by nitrifying an ammonium plume, and subsequently (ii) to deliver a reductant (e.g., ethanol) to denitrify nitrite/nitrate to nitrogen gas. (C) Primary soil column result showing atrazine attenuation (decrease to near zero) immediately when oxygen was supplied in air at mid-way in the soil column (after [50]). (D) Primary soil column result showing ammonium being transformed during addition of oxygen and then nitrite/nitrate being degraded on addition of ethanol via the polymer mats (after [96]).

For the ammonium plume PRB example (Fig. 3B), a dual sequential treatment strategy was adopted to first induce nitrification of the ammonium plume by use of air addition through polymer mats followed by denitrification to nitrogen gas via addition of a reductant (e.g., carbon amendment or hydrogen gas) again via use of the polymer tubing mats. In Fig. 3D, primary data from the soil column investigations shows the successful sequential transformation with the second woven polymer tube mat supplying ethanol to denitrify the nitrite/nitrate generated earlier in the column. In laboratory columns, half-life nitrification rates (ammonium to nitrite/nitrate) were 0.07–0.25 days. Adding hydrogen sequentially to induce denitrification of the created nitrite/nitrate gave a half-life of 3.5 days, whilst using ethanol gave a half-life of 0.12–0.34 days [96]. The technique was implemented in the field in a funnel and gate PRB design using ethanol as the reductant [133].

A range of gas injection strategies have been adopted to remediate contaminants in the subsurface. Air sparging strategies (Fig. 4A, Ref. [112, 134]) were developed largely to remove volatiles such as BTEX or chlorinated volatiles but can also induce biodegradation. During air sparging of a dissolved gasoline plume, Johnstonet al. (1998) [112] found that aerobic biodegradation was occurring, but during the period of largest mass removal, biodegradation was occurring at a rate an order of magnitude less than volatilization rates estimated by capture of extracted volatiles. In the field study, air sparging started on day 10 but was stopped for 4-h periods on days 10, 11 and 13 to evaluate changes in dissolved oxygen and volatile organic compounds (VOCs) in groundwater. Data from the multi-depth monitoring location denoted DP2 (Fig. 4C), 1 m from the sparge well, show persistent reduction in benzene concentrations over time at shallower depths — say 3.5 and 4.5 m depths. Reductions in concentrations were observed deeper, but with significant rebound in concentrations when injection was halted. These depths appear to be towards the outer edge of the zone of influence for the sparge well in this sandy aquifer. Such data were used to both calculate biodegradation rates but also to optimize the spacing of sparge points across the gasoline groundwater plume at the site to effect remediation.

Fig. 4.

(A) Schematic of an air sparging system to remediate contaminated groundwater, showing the induced dual processes of volatilisation and biodegradation (where oxygen is utilized). (B) Schematic of bioventing remediation, showing air injection above the water table, drawdown of the water table to expose the zone of contamination to gaseous oxygen and nutrient addition through injection points. (C) Benzene concentrations during air sparging at multiple depths at monitoring well 1 m from the sparge well. Note that air sparging commenced on day 10 (after [112]). (D) Diesel NAPL in soil cores with depth below ground at two locations (Y8: where no nutrients were added; Y11: where nutrients were periodically added) from prior to 14 months of bioventing (yellow data), and at the end (orange data) (after [134]).

To maximize biodegradation and minimize the need to capture volatile emissions in extracted soil gas (Fig. 4A), biosparging strategies were developed using periodic gas sparging compared to continuous or high pressure/flows [116] avoiding the need to capture volatilized compounds. A similar approach was taken by Davis et al. (2009) [51] who tracked the production of chloride, bicarbonate and other indicators to quantify the aerobic biodegradation of 1,2 DCA and vinyl chloride in groundwater induced by air injection into a biosparging curtain across the groundwater flow direction. Other studies have investigated sparging/biosparging using reductant gases such as hydrogen to bioremediate chlorinated hydrocarbons [133, 135] and perchlorate and nitrate plumes [136]. An overview of air sparging and biosparging was published in Suthersan et al. (2017) [137].

Bioventing is the injection of gaseous amendments (usually air) above the water table (c.f., air sparging which is below the water table), often used to target contaminants like petroleum fuels residing at or across the zone of water table fluctuation (Fig. 4B) [138]. It can be well suited to removing significant mass from the source zone that might otherwise slowly dissolve and continue to be a source of soluble contaminants into a groundwater plume. Sometimes the water table is drawn down via pumping to expose the contaminated horizon to an injected gas phase, and nutrients might be added to increase biodegradation rates as depicted in Fig. 4B. Such a trial was implemented at a refinery site [139] for weathered diesel, and over 14 months biodegradation rates were calculated from oxygen consumption (amongst other measures) [134, 140]. Biodegradation rates increased over the trial from 5–15 mg/kg-soil/day up to 30–90 mg/kg-soil/day. Key outcomes are shown in Fig. 4D. Near complete diesel NAPL mass removal was achieved for the depth profile where nutrients (ammonium nitrate and some phosphorus) were added periodically (Y11). Drawdown of the water table and targeted nutrient delivery to the zone of contamination were seen as critical to successfully increasing biodegradation rates and mass removal at Y11. Significant mass removal also occurred where no nutrients were added (Y8) (Fig. 4D), but complete diesel mass removal over the time of bioventing did not occur at Y8.

Compilations of bioremediation performance data at field sites across all contaminants discussed here are not available, however assessments have been made for chlorinated solvent and petroleum hydrocarbons. Performance data across 117 chlorinated hydrocarbon anaerobic bioremediation projects showed a median reduction in source zone primary contaminant concentrations of 91% after the bioremediation project was implemented [141]. However, only 7% of the projects achieving the United States drinking water criteria. A similar compilation for petroleum hydrocarbon sites showed a median reduction in source concentration post bioremediation of just less than 90% with 11% of the projects achieving the U.S. drinking water criteria [142]. There are likely several caveats on the estimated means; one of which may be the dependence on mass removal effectiveness of prior site characterization efforts and the efficacy of the methods adopted to deliver amendments to contamination targets in the subsurface; another caveat may be that improved technologies, knowledge and approaches may have been available at more recent bioremediated sites compared to sites undergoing bioremediation initially. Never-the-less, improvements in bioremediation are needed to meet closure criteria in an efficient way and in targeting risk drivers.

In Section 2 we discussed features of groundwater systems and their role in biodegradation of a contaminant plume. The longevity and extent of groundwater plumes relies also on the ‘strength’ of the source of the plume [143] – i.e., both the contaminant concentration and the net flux of dissolved components at the source entering the plume. There are increasing efforts to quantify the source strengths of plumes, and to understand what mass depletion from source zones might be required to reduce concentrations below regulatory criteria [144], and fluxes to a level that would enable biodegradation and other attenuation mechanisms to accelerate the shrinkage of plumes. As mentioned earlier, for petroleum hydrocarbons, focus has moved from MNA of plumes to NSZD which seeks to quantify all groundwater and vadose zone biodegradation and mass loss including volatile losses [145, 146]. Longer term trends over time in NSZD whole-of-source biodegradation rates, and overall mass losses are being investigated [139]. Because of the low apparent biodegradability of PFAS many investigations are underway to quantify [19] and limit [147] the leaching of PFAS from source soils and vadose zones, to reduce the likelihood of exceeding low regulatory compliance criteria in groundwater and receiving water bodies. Indeed, even after primary plume sources are removed, secondary sources may lead to back-diffusion from tight, low-conductivity, or less mobile parts of aquifers and soils which can continue to feed groundwater plumes over long periods despite best efforts to remediate [148, 149]. Some remedial strategies have also targeted direct removal of source zone mass below the water table to seek to limit groundwater plume migration and length [150, 151] and to enable accelerated shrinkage of plumes via biodegradation processes [152]. Defining the level of effort required in remediating whole plumes to regulatory acceptance criteria requires quantitative integration of all key bio-geophysical and chemical processes within soil and plume source zones and throughout a groundwater plume’s extent (as per Figs. 1,2), and where needed inclusive of interactions and contact with receptors (e.g., ecosystems or groundwater extraction from wells). We provide an overview of some integrative modelling efforts in Section 6.3.

Models require input parameters such as the functional form of biodegradation rate kinetics. Biodegradation rates may change due to a range of factors, inclusive of the type and concentration of the parent compound being degraded, the growth, decay, and transition in suitable microbial populations [36, 153], the creation of degradation products, limitations due to nutrient or other carbon sources, competing contaminants when mixtures are present, temperature effects [154] or inhibitions due to toxicity of some contaminants. Early laboratory batch and microcosm studies simply graphed and analyzed concentration changes over time [155] to formalize kinetic functional forms [156]. A linear decreasing concentration over time implied a constant biodegradation rate unaffected over time and is termed zero-order kinetics [134, 140]. An exponentially decreasing concentration curve over time, was often fitted to a first-order kinetic rate meaning the rate was proportional to the concentration of the parent compound [157]. Second or nth order reaction kinetics have also been explored to fit observed data [158]. It is noted that concentration vs. time data from monitoring wells is not a biodegradation rate of dissolved constituents but more a reflection of source zone attenuation rates, and that concentration vs. distance rates reflect bulk attenuation coefficients that include primarily biodegradation with some dispersion [159].

Monod rate kinetics [160] are often used as they set a maximum asymptotic rate of reaction that reflects the metabolic capabilities of the microbial population. It is also a formulation that transitions between approximately first-order kinetics for low concentrations to zero-order rate kinetics for high concentrations. A number of studies have compared zero-order, first-order and Monod models in representing observed data [161, 162]. Chambon et al. (2013) [162] found when modelling reductive de-chlorination that zero-order and Monod kinetic parameters varied over a wide range, and that experimental microbial data were scarce. They surmised that the wide range of parameters was because reductive de-chlorination involved many microbial populations and reactions (e.g., fermentative, methanogenic, iron reducing) that were not accommodated by simple functional forms of the rate kinetics.

Monod kinetics are often used in groundwater biodegradation models and sometimes multiplications of many Monod kinetic functional forms are used to represent limitations and rate controls due to additional chemical species that may be needed in biodegradation reactions, or indeed those that might inhibit biodegradation [163, 164]. Rifai and Bedient (1990) [165] compared a dual substrate Monod kinetic relationship to calculate biodegradation to that of an instantaneous biodegradation rate where the rate is limited by transport processes (perhaps an electron acceptor or donor transported to the zone of contamination). They found that whilst the Monod model might be more accurate, it also required estimation of additional parameters. Borden et al. (1986) [166, 167] introduced the concept of instantaneous reaction kinetics for groundwater plumes whereby all oxygen and hydrocarbon would be largely instantly consumed by microbial controlled biodegradation only limited by the transport and availability of oxygen and hydrocarbons. This was similarly proposed for oxygen and hydrocarbon vapor biodegradation in vadose zones [168].

Kinetic rate formulations are simplified representations of biodegradation processes. Accommodating all the pathways and sub-processes responsible for contaminant biodegradation in kinetic models would lead to overly complex systems with a large number of unresolved parameters that would require estimation or fitting. Often the underlying detail of multiple microbial groups contributing to many sub-reaction pathways are unable to be accommodated [162]. Assumptions regarding kinetics and which key chemical species provide the appropriate geochemistry suitable for biodegradation to proceed are necessarily simplified in kinetic models.

The range of models that seek to simulate biodegradation processes in groundwater is vast [163, 166, 169]. Models that integrate groundwater flow aspects, geochemical reactions and adequately represents microbial biomass responses are necessary to constrain the potential natural rates of attenuation that may be estimated for aquifers, and indeed to evaluate designs for accelerating bioremediation. Simplified models have been developed (e.g., BIOSCREEN and BIOCHLOR [170, 171]) for approximating plume dimensions. These commonly assume groundwater flow rates and a degradation half-life. The dispersion coefficient is also a key parameter as a measure of dilution and mixing. However, such models can easily provide unrealistic estimates of biodegradation and hence plume length when key coupled processes are not incorporated and when not calibrated to actual site data.

Clement et al. (1998) [172] developed RT3D which described discrete transformation pathways for contaminants with known degradation pathways and applied it to a range of contaminants [173]. PHT3D [43, 174] was one of the first models to link multicomponent groundwater plume reactive transport modules to geochemical and redox controls on contaminant concentration changes during biodegradation. The framework also ensured limited numerical dispersion errors so as not to incorrectly overmagnify mixing and hence biodegradation within and on the fringes of plumes [44]. Linking the geochemical equilibration code PHREEQC provided additional indicators of biodegradation such as inorganic chemistry changes, but also assisted with constraining model simulations to more realistic site geochemical conditions [175, 176]. To additionally simulate microbial population mobility and attachment, Wang and Corapcioglu (2002) [177] modelled bioaugmentation using modified Monod kinetics and a microcolony concept to investigate the relative effects on the transport and biodegradation of an organic contaminant. Increasingly the dynamics of microbial populations and their role in biodegradation is being embedded into reactive transport models [178, 179]. New multi-omics tools promise vast data sets; accommodating it meaningfully in reactive transport models to advance the understanding of biodegradation processes is a future challenge.

Comparison of model outputs to laboratory or field data is key to determining if models accommodate most of the underlying processes being investigated. When fitted adequately to field or laboratory data [164, 180, 181, 182, 183, 184, 185] the model provides a diagnostic tool to assess scenarios, to predict future trends and contaminant plume fate — but also provides a platform for assessing potential remedial options and designs to accelerate and optimize bioremediation of groundwater plumes [186, 187, 188]. When not calibrated or constrained by data, model predictions can be uncertain, and use of such models may pose a risk in assessing underlying processes incorrectly. Appropriately calibrated models are not only becoming more powerful and representative of biodegradation processes in groundwater — but are essential to addressing the challenges of managing the risk and remediation of contaminant plumes in groundwater, and groundwater quality more generally [183].