In Situ Chemical Oxidation of Petroleum Hydrocarbons

Abstract

In situ chemical oxidation (ISCO) is a mature treatment technology that involves the delivery of a chemical oxidant into a target treatment zone (TTZ) to destroy petroleum hydrocarbon (PHC) compounds, and thereby reduce risk to human health and the environment. Commonly used chemical oxidants include hydrogen peroxide, sodium percarbonate, ozone, sodium or potassium permanganate, and sodium or potassium persulfate. All these oxidants can degrade environmentally relevant PHCs except that permanganate is non-reactive toward benzene. Ozone is delivered into the TTZ as a gas while the other oxidants are typically delivered as a concentrated liquid. ISCO should be considered part of integrated remediation strategy and not used in isolation. This chapter provides a reader with an introduction to key aspects of ISCO that are relevant to applications at sites contaminated with PHCs. Following a discussion of the TTZ, it then examines the fundamentals of the common oxidants used. Next the interaction of chemical oxidants with aquifer materials is explored, and a description of relevant transport considerations is provided. This is followed with an overview of methods that can be used to deliver an oxidant to treat PHCs in a TTZ. This chapter closes with a summary of the important takeaway messages.

15.1 Introduction

An improved awareness of the health risk associated with petroleum hydrocarbons (PHCs), and an increased scientific understanding of the underlying processes related to their fate and transport over the past 30 years has fostered the growth and introduction of a number of remedial technologies. These technologies range from passive methods (e.g., monitored natural attenuation, and permeable reactive barriers) to more active methods [e.g., soil vapor extraction (SVE), in situ air sparging (IAS), and pump-and-treat (P&T)]. Technology selection depends on a number of factors including site characteristics (e.g., type and extent of PHC contamination, and hydrogeological conditions), and is influenced by robustness and economic viability (FRTR 2022).

In situ chemical oxidation (ISCO) is based on an extension of the already well-established water and wastewater treatment processes wherein chemical oxidants are added to water to destroy target contaminants and thereby improve water quality (Colthurst and Singer 1982; Cleasby et al. 1964; Houston 1918). Similarly, ISCO involves the delivery of a strong oxidizing reagent into the subsurface to transform PHC compounds into less harmful intermediates or end-products such as carbon dioxide (CO₂), and thereby reduce potential risk to public health and the ecosystem. While complete mineralization is a worthwhile objective, partial chemical oxidation where long-chain PHCs are converted into less-complex, more water-soluble, and more biodegradable compounds or intermediates is a beneficial objective (Karpenko et al. 2009).

ISCO is potentially able to operate over a wide range of conditions, and provides economic advantages over other conventional technologies due, in part, to its aggressive nature (ITRC 2005). There are a variety of chemical oxidants that are currently being used in ISCO applications, including hydrogen peroxide (H₂O₂), sodium percarbonate (Na₂CO₃·1.5H₂O₂), ozone (O₃), sodium or potassium permanganate (NaMnO₄, KMnO₄), and sodium or potassium persulfate (Na₂S₂O₈, K₂S₂O₈).

The design of an ISCO system involves a comprehensive understanding of the underlying physical and chemical processes, and in general, requires an oxidant to have a high rate and extent of reactivity toward target PHCs, and a high stability and persistence in the subsurface. These criteria serve to establish the usefulness, effectiveness, and efficiency of an oxidant.

Since the initial application of ISCO in 1984 (Brown et al. 1986), there has been substantial research and development activities that have generated a tremendous amount of published literature dealing with oxidant chemistry, oxidant interactions with aquifer materials, transport processes, delivery methods, and combined or integrated treatment systems (e.g., see Siegrist et al. 2011; USEPA 2006; ITRC 2005). In addition to the published literature, the number of field applications has grown throughout the world to a point where ISCO is considered a mature technology and has been added to the remedial toolbox.

The aim of this chapter is to provide a reader with an introduction to key aspects of ISCO that are relevant to applications at sites contaminated with PHCs. It begins with a discussion of the target treatment zone, and then examines the fundamentals of the most commonly used oxidants. Next the interaction of chemical oxidants with aquifer materials is explored, and a description of relevant transport considerations is provided. This is followed with an overview of methods that can be used to deliver an oxidant to treat PHCs in the subsurface. Finally, the chapter closes with a summary of the important takeaway messages. There are numerous ISCO case studies available in the literature for an interested reader to consult (see USEPA 2022).

15.2 The Target Treatment Zone

Typically, ISCO should not be considered until all attempts to remove mobile light non-aqueous phase liquid (LNAPL) from the subsurface have been exhausted. The mass of oxidant required to oxidize mobile LNAPL, at most sites, will not be economically feasible. Since chemical oxidants are not able to mix with the LNAPL (i.e., they are not miscible), all the important chemical oxidation reactions occur in the aqueous phase. Thus, mass associated with the LNAPL must dissolve into the aqueous phase for it to be accessible to the oxidant. Mass transfer from the LNAPL to the aqueous phase is controlled by a mass transfer rate coefficient and the aqueous solubility of the PHC compound of interest. For some PHCs and LNAPL architectures, the mass transfer rate is limiting rather than the chemical oxidation reaction rate. Therefore, the removal of mobile LNAPL using recovery methods such as skimming, vacuum enhanced recovery, and multiphase extraction (ITRC 2018) is commonly recommended before any ISCO application. ISCO should not be viewed as a stand-alone treatment technology, but rather as integral part of an integrated treatment approach.

After the majority of the mobile LNAPL has been removed, ISCO can be used to treat the remaining LNAPL source zone and the associated dissolved phase plume. Figure 15.1 conceptually illustrates a residual LNAPL source zone and dissolved phase plume in profile and plan views. The objective of an effective ISCO application is to create an in situ reactive zone (IRZ) where target PHC compounds are oxidized to less toxic end products. The design of an IRZ requires that the delivered oxidants can react with the target PHCs of concern at a sufficient rate to overcome competing reactions, and that enough oxidant mass is distributed throughout the IRZ to support the demand (Suthersan and Payne 2005). ISCO treatment objectives need to be clearly stated so that expectations are non-ambiguous. The target treatment zone (TTZ) may be the residual LNAPL in the unsaturated zone, the entrapped LNAPL mass in the saturated capillary fringe and the saturated groundwater zone (smear zone), and/or the dissolved phase plume. While the IRZ design principles for each of these TTZs are similar, the way in which they are implemented is different. Site characterization data must be carefully reviewed to determine if, and where, LNAPL is likely present. Persistent dissolved PHCs plumes typically suggest the presence of LNAPL, and the TTZ should encompass those areas of the site where LNAPL sources are suspected to be present.

Fig. 15.1

A conceptual illustration of an LNAPL source zone and dissolved phase plume in a profile view, and b plan view. The source zone is presented as a smear zone in the capillary fringe and saturated groundwater zone with a core of higher LNAPL saturation. The dissolved phase plume is also shown with a higher concentration core

In addition to PHC mass distribution, consideration must also be given to the delivery of a liquid or gas oxidant into the TTZ. A TTZ in unconsolidated porous media (e.g., a mixture of sand, silt, and clay) is considered suitable for ISCO if the average saturated hydraulic conductivity (K) is > 10^–6 m/s. The spatial variation or heterogeneity of K in the TTZ is also important with oxidant delivery into highly heterogeneous TTZs being problematic (e.g., variations in K of > 10³ or 10⁴ m/s).

15.3 Oxidant Fundamentals

The kernel of the any ISCO system is its capability to quickly react with the target PHC compounds of interest, and degrade them into innocuous end-products such as CO₂ and water. While complete mineralization of the target PHCs is the ultimate goal, often there are many reaction steps required to reach this goal and numerous intermediates or oxidation by-products will be produced. Usually these by-products are more water-soluble, less toxic, and susceptible to biodegradation. For example, a chemical oxidation pathway for benzene involves the production of phenol which is then transformed to catechol (a benzenediol) (Fu et al. 2017). While catechol can also be chemically oxidized, it is readily biodegraded since it is also formed during the aerobic degradation of benzene (e.g., Yu et al. 2001).

For the commonly used oxidants discussed in this chapter, Table 15.1 indicates their ability to degrade benzene, ethylbenzene, toluene, and xylenes (BTEX), mixtures of PHCs as represented by total petroleum hydrocarbons (TPH), and polycyclic aromatic hydrocarbons (PAHs) (e.g., naphthalene, fluorene, and anthracene). As shown, all oxidants are amenable to degrade the target PHCs listed except that permanganate is essentially non-reactive toward benzene. TPH is a bulk measure of the petroleum-based aromatic and aliphatic hydrocarbons found in gasoline, diesel, and oil range organics [e.g., see US EPA Method 8015D (SW-846)]. While Table 15.1 indicates that TPH can be degraded by the oxidants listed, this does not imply that complete mineralization of the total mass associated with TPH is possible, rather that TPH has been observed to decrease following exposure to one of the oxidants. Similarly, only some individual PAHs have been reported to be susceptible to oxidation and thus the total PAH sum, which is often used as a treatment metric, will decrease following exposure to an oxidant.

Table 15.1 General ability of the most commonly used oxidants to degrade BTEX compounds, TPH, and PAHs

Often the standard reduction potential (E^o) is used to compare the potential of various oxidant species. The standard reduction potential is the potential in volts (V) generated by a half-reaction relative to the standard hydrogen electrode (SHE) under standard conditions (25 °C, 1 atm, concentration of 1 M). The larger the potential the greater desire the species has to acquire electrons, or the more the species tends to be reduced (Snoeyink and Jenkins 1980). While E^o values are informative, they do not provide information on reaction stoichiometry or kinetics for a specific PHC.

The rate of reaction is an important consideration for ISCO treatment effectiveness, and hence numerous reaction rate coefficients have been reported in the literature for a variety of PHCs, oxidant systems, and reaction conditions. Because many of these reactions occur in the aqueous phase, the chemistry of the oxidants and their interaction with PHC compounds have been the subject of intensive study. Typically, the second-order mass action law shown in Eq. (15.1) is used to represent the aqueous phase reaction kinetics for a PHC compound exposed to an oxidant (Forsey et al. 2010; Levenspiel 1999)

$$\frac{{{\text{d}}[C_{{{\text{PHC}}}} ]}}{{{\text{d}}t}} = \text{-} k^{\prime\prime}[C_{{{\text{PHC}}}} ][C_{{{\text{ox}}}} ]$$

(15.1)

where [C_PHC] is the molar concentration of the PHC compound (mol/L), [C_ox] is the molar concentration of the oxidant (mol/L), and $k^{\prime\prime}$ is the second-order rate coefficient (L/mol s or M^-1 s^-1). Tratnyek et al. (2019) complied a dataset of $k^{\prime\prime}$ values that were available as of 2005 which include data for a select number of PHC compounds. Treatability studies which are used to generate kinetic data are often limited since they require simplifying assumptions that may ignore changing reaction conditions (pH shifts and radical complexities), and generation and consumption of intermediates and by-products. In these cases, a pseudo first-order mass action law (Eq. 15.2) is often used to fit the kinetic data as given by

$$\frac{{{\text{d}}[C_{{{\text{PHC}}}} ]}}{{{\text{d}}t}} = \text{-} k^{\prime}_{{{\text{obs}}}} [C_{{{\text{PHC}}}} ]$$

(15.2)

where $k^{\prime}_{{{\text{obs}}}}$ is the observed first-order rate coefficient (s^-1) associated with the degradation of the parent or target PHC compound. In addition to the concentration of the reactants (Eq. 15.1), the in situ degradation rate of a specific PHC compound is dependent on many variables that also play a role including temperature, pH, interaction with aquifer solids or materials, and the concentration of catalysts, reaction by-products, natural organic matter, and scavengers.

The following sub-sections provide a high-level overview of the commonly used oxidants (hydrogen peroxide, percarbonate, ozone, permanganate, and persulfate) since a basic understanding of the reactions involved is required. For some oxidants the suite of reactive species is too complicated, so a simplified perspective is discussed. Also provided is a short synopsis of the capability of each oxidant to degrade target PHCs based on laboratory studies. By no means is the description of each oxidant presented in this section complete, but it is considered to be sufficient to appreciate their unique qualities. For additional insight, an interested reader is encouraged to consult Siegrist et al. (2011, 2014), USEPA (2006) and ITRC (2005), and the vast list of references contained within.

15.3.1 Hydrogen Peroxide

Hydrogen peroxide (H₂O₂) is a strong liquid oxidant that is used in many industrial applications, and in water treatment processes. Hydrogen peroxide is often applied in combination with a catalyst or activator, and the term catalyzed hydrogen peroxide (CHP) has been used to refer to these hydrogen peroxide systems. When catalyzed or activated in water, hydrogen peroxide generates a wide variety of free radicals and other reactive species. A radical or free radical is a cluster of atoms that contains an unpaired electron. This configuration is extremely unstable, and hence radicals quickly react to achieve a stable configuration. Reactive species formed in a CHP system can include both oxidants and reductants.

As an alternative to the liquid form of H₂O₂, various forms of solid peroxides (e.g., percarbonate, and peroxide) have been used. Solid forms of H₂O₂ are more easily stored and transported than the liquid form of H₂O₂. At ambient temperature, sodium percarbonate (Na₂CO₃·1.5H₂O₂) will yield H₂O₂, sodium (Na⁺), and carbonate (${\text{CO}}_{{3}}^{{2 \text{-} }}$) when dissolved in water as given by Eq. (15.3):

$${\text{2Na}}_{{2}} {\text{CO}}_{{3}} {\cdot} {\text{3H}}_{{2}} {\text{O}}_{{2}} \to {\text{3H}}_{{2}} {\text{O}}_{{2}} + {\text{4Na}}^{ + } + {\text{2CO}}_{{3}}^{{2 \text{-} }}$$

(15.3)

Although calcium peroxide (CaO₂, or magnesium peroxide, MgO₂) is usually considered an oxygen releasing compound used to promote aerobic biodegradation, it produces H₂O₂ when dissolved in water (Eq. 15.4)

$${\text{CaO}}_{{2}} + 2{\text{H}}_{{2}} {\text{O}} \to {\text{H}}_{{2}} {\text{O}}_{{2}} + {\text{Ca}}^{{2 + }} + 2{\text{OH}}^{ \text{-} }$$

(15.4)

By itself, H₂O₂ can directly react with a PHC compound through a direct two-electron transfer (Eq. 15.5a). While this reaction has a high standard reduction potential (E^o = 1.8 V) the rate of these reactions is deemed slow and not relevant for ISCO applications (Watts and Teel 2005).

The chemistry of the CHP system grew from the well-established Fenton’s reagent (Fenton 1894) in which ferrous iron (Fe²⁺) salts were used to catalyze a dilute solution of H₂O₂ at a pH between 3 and 5 to yield hydroxyl radicals (${}^{ \cdot }{\text{OH}}$), ferric iron (Fe³⁺), and hydroxyl ions (Eq. 15.5b). The generated ferric iron then reacts with H₂O₂ or, with another produced radical, the hydroperoxyl radical (${\text{HO}}_{2}^{ \cdot }$, E^o = 1.7 V) to produce Fe²⁺ (Eqs. 15.5c and 15.5d). The reactions (Eqs. 15.5b–15.5d) will essentially deplete the available H₂O₂ in a self-destructing manner. The hydroxyl radical is a non-selective oxidant that has an unpaired electron making it highly reactive (E^o = 2.8 V) with a wide variety of PHC compounds.

$${\text{H}}_{{2}} {\text{O}}_{{2}} + {\text{2H}}^{ + } + {\text{2e}}^{\text{-}} \to {\text{2H}}_{{2}} {\text{O}}$$

(15.5a)

$${\text{H}}_{{2}} {\text{O}}_{{2}} + {\text{Fe}}^{2 + } \to {\text{OH}}^{ \text{-} } + {\text{Fe}}^{3 + } + {}^{ \cdot }{\text{OH}}$$

(15.5b)

$${\text{Fe}}^{3 + } + \;{\text{H}}_{{2}} {\text{O}}_{{2}} \to {\text{Fe}}^{2 + } + {\text{H}}^{ + } + \;{\text{HO}}_{{2}}^{ \cdot }$$

(15.5c)

$${\text{Fe}}^{3 + } + {\text{HO}}_{{2}}^{ \cdot } \to {\text{Fe}}^{2 + } + {\text{H}}^{ + } + {\text{O}}_{{2}}$$

(15.5d)

$${\text{HO}}_{{2}}^{ \cdot } \leftrightarrow {\text{O}}_{{2}}^{{ \cdot { \text{-} }}} + {\text{H}}^{ + }$$

(15.5e)

$${\text{Fe}}^{2 + } + {\text{HO}}_{{2}}^{ \cdot } \to {\text{Fe}}^{3 + } + {\text{HO}}_{{2}}^{ \text{-} }$$

(15.5f)

$${}^{ \cdot }{\text{OH}} + {\text{Fe}}^{2 + } \to {\text{Fe}}^{3 + } + {\text{OH}}^{ \text{-} }$$

(15.5g)

$${\text{H}}_{{2}} {\text{O}}_{{2}} + {}^{ \cdot }{\text{OH}} \to {\text{H}}_{{2}} {\text{O}} + {\text{HO}}_{{2}}^{ \cdot }$$

(15.5h)

As the concentration of H₂O₂ is increased, several other important radicals and ions are generated including the superoxide radical (${\text{O}}_{{2}}^{{ \cdot {\text{-}}}}$, Eq. (15.5e); E^o = - 2.4 V) and the hydroperoxide anion [${\text{HO}}_{{2}}^{ \text{-} }$, Eq. (15.5f); E^o = - 0.9 V]. In contrast to the hydroxyl and hydroperoxide radicals which are oxidants, the superoxide radical and hydroperoxide anion are reductants providing a reduction degradation pathway for some PHC compounds (Furman et al. 2009; Smith et al. 2004). The reactions given by Eqs. (15.5b)–(15.5h) occur nearly simultaneously and will continue until one of the reactants becomes limiting which is normally H₂O₂.

One of the important by-products of CHP reactions is the generation of large quantities of dissolved oxygen (O₂) that is available in the TTZ after H₂O₂ has been depleted. Since O₂ is an electron acceptor for the aerobic biodegradation of many PHC compounds, its presence may lead to further degradation of target PHCs and intermediates of chemical oxidation. However, O₂ evolution in some cases may result in the mobilization of volatile PHCs leading to unacceptable exposure pathways. CHP reactions are exothermic, and therefore they can result in significant temperature increases within the TTZ. Heat production is dependent on the concentration of H₂O₂ used and the rate of reaction. This excess heat and perhaps steam is a health and safety hazard which may need to be addressed.

To enable important CHP reactions to occur in the TTZ, ferrous sulfate (FeSO₄) or other Fe²⁺ salts are typically co-injected with H₂O₂ at concentrations significantly above background levels. In porous media, both Fe²⁺ and Fe³⁺ are susceptible to numerous reactions (e.g., precipitation, complexation) which eliminate them from the CHP reactions. More importantly perhaps, is that the solubility of Fe²⁺ at typical groundwater pH levels (7–9) is limiting. To overcome these issues, stabilizers are typically used (Checa-Fernandez et al. 2021). Phosphate was one of the initial stabilizing compounds used; however, since it is a strong complexing agent, it restricted the availability of Fe²⁺. Watts et al. (2007) suggested that the most effective stabilizers are citrate, malonate, and phytic acids.

Numerous non-productive or scavenging reactions compete with the target PHCs for the radicals generated in a CHP system. The most common scavenger anions prevalent in groundwater systems are carbonate, bicarbonate, chloride, and sulfate. These anions react with ${}^{ \cdot }{\text{OH}}$ to form the bicarbonate radical (Eq. 15.6a), the carbonate radical (Eq. 15.6b), the hypochlorite radical (Eq. 15.6c), and the sulfate radical (Eq. 15.6d). While these scavenging reactions are generally considered unproductive, evidence suggests that carbonate radicals are capable of oxidizing some aromatic compounds (e.g., benzene) (Umschlag and Herrmann 1999), and the sulfate radical has a high reactivity toward PHC compounds (see Sect. 15.3.4).

$${}^{ \cdot }{\text{OH + HCO}}_{{3}}^{\text{-}} \to {\text{OH}}^{\text{-}} {\text{ + HCO}}_{{3}}^{ \cdot }$$

(15.6a)

$${}^{ \cdot }{\text{OH + CO}}_{{3}}^{{2 \text{-} }} \to {\text{OH}}^{\text{-}} {\text{ + CO}}_{{3}}^{{ \cdot {\text{-}}}}$$

(15.6b)

$${}^{ \cdot }{\text{OH + Cl}}^{\text{-}} \to {\text{ClOH}}^{{ \cdot {\text{-}}}}$$

(15.6c)

$${}^{ \cdot }{\text{OH + SO}}_{{4}}^{{2 \text{-} }} \to {\text{OH}}^{\text{-}} {\text{ + SO}}_{{4}}^{{ \cdot {\text{-}}}}$$

(15.6d)

In some instances, the presence of minerals associated with aquifer materials may act like a catalyst similar to the addition of Fe²⁺ discussed above. The minerals of relevance to ISCO are those associated with Fe or Mn since they are typically observed at the highest concentration in aquifer materials. Goethite (α-FeOOH), hematite (Fe₂O₃), and pyrolusite (β-MnO₂) have been shown to catalyze H₂O₂ (e.g., Kwan and Voelker 2003, Teel et al. 2001). In contrast to the homogeneous reaction between H₂O₂ and dissolved Fe²⁺, the reaction between H₂O₂ and minerals is heterogeneous since it occurs at the mineral surface.

Prior to delivery, percarbonate is typically mixed with a combination of silicates and FeSO₄. When added to water, an alkaline pH version of the CHP system will be generated. In this base-catalyzed peroxide system, the predominant species that is generated is the hydroperoxyl radical (${\text{HO}}_{2}^{ \cdot }$). Na₂CO₃ released will form carbonate (${\text{CO}}_{{3}}^{{2 \text{-} }}$) and bicarbonate (${\text{HCO}}_{{3}}^{\text{-}}$) which will scavenge radicals [see Eqs. (15.6a) and (15.6b)]. Ma et al. (2020) reported that the lifetime of percarbonate is relatively short which limits its availability in the subsurface.

CHP systems have been demonstrated in many laboratory studies to treat PHCs with success. For example, Watts and Dilly (1996) investigated the ability of phosphate stabilized H₂O₂ catalyzed by six different iron compounds to oxidize diesel sorbed to soils (1000 mg/kg). All iron catalysts promoted > 70% degradation in < 1 h. Xu et al. (2006) used H₂O₂ and a mixture of ferrous salts to oxidize diesel spiked soils, and reported that the mass removed (as high at 93%) was dependent on the concentration and volume of H₂O₂ as well as the soil composition. Usman et al. (2012) investigated the capability of magnetite (Fe₃O₄) catalyzed H₂O₂ to treat soil spiked with fresh crude oil and weathered crude oil. The results showed that 84% of the weathered crude oil and 92% of the crude oil was degraded in one week. Surprisingly, these authors also report that no by-products were observed and thus complete degradation of PHCs was achieved. Aromatic compounds (i.e., BTEX) have been found to be easily degraded by ${}^{ \cdot }{\text{OH}}$. Table 15.2 provides the second-order reaction rate coefficients for the reaction between $^{ \cdot } {\text{OH}}$ and BTEX compounds (Buxton et al. 1988).

Table 15.2 Second-order reaction rate coefficients for the oxidation of BTEX compounds by hydroxyl radicals (from Buxton et al. 1988)

Since PAHs are comprised of multiple aromatic benzene rings, oxidation of PAHs by CHP systems is possible. Lundstedt et al. (2006) reported that some PAHs (anthracene, benzo[a]pyrene, and perylene) were extensively degraded compared to other PAHs with fewer or equal number of benzene rings during CHP treatment of contaminated soil collected from a former gasworks site indicating selectively. Nevertheless, considering the complex structure of the higher ringed PAHs, it would be expected that numerous ${}^{ \cdot }{\text{OH}}$ reactions would be required to fully mineralize the parent PAH and, in this process, various intermediates would be generated (Lundstedt et al. 2006).

15.3.2 Ozone

Ozone (O₃) is a highly reactive and unstable gas, and must be generated at the point of use with an ozone generator. Onsite O₃ generators that use an air supply can produce an O₃ concentration of 1–2% by volume, so the bulk of the injected gas is atmospheric air. This gas can be delivered directly into the unsaturated zone or sparged into the saturated zone. In the latter case, the distribution of O₃ will be controlled by the same processes that control the air distribution during in situ air sparging (Tomlinson et al. 2003; Thomson and Johnson 2000). Figure 15.2 provides a schematic representation of the injection of ozone gas below a LNAPL source TTZ, and the observed spatial distribution of gas release locations at the water table. The nature of the discrete gas pathways within the TTZ suggests that treatment will be limited by mass transfer of O₃ to the aqueous phase.

Fig. 15.2

a Schematic representation of the injection of ozone gas below a LNAPL TTZ. Near the injection screen there will be a region of elevated gas content; however, due to flow instabilities a network of isolated gas channels will form and the gas will migrate vertically upward to escape into the unsaturated zone. b Spatial distribution of non-uniform gas release locations at the water table during steady-state air sparging (from Tomlinson et al. 2003). Each observed air release location is indicated by a circle with a radius corresponding to the relative intensity of the air release at that location (smallest is the least intense). Grid spacing in northerly and easterly directions is 1 m

Oxidation reactions can occur both in the gas and aqueous phases. Gas phase reactions will occur when PHC compounds have volatilized into the gas phase and react with O₃. Aqueous phase reactions require O₃ to partition into the groundwater which is driven by the O₃ concentration gradient between the gas and aqueous phases; however, the aqueous solubility of O₃ is low (~28 mg/L at 10 °C from a 5% O₃ gas concentration) which limits the mass of O₃ available in the aqueous phase. This limitation is compounded by the discrete gas channels that develop in the saturated zone as discussed above. Once in the aqueous phase, a number of decomposition processes involving organic and inorganic reactants exert a constant demand for O₃.

Tomiyasu et al. (1985) proposed the ozone decomposition pathway given by Eqs. (15.7a–15.7i). This pathway begins with a one-electron transfer from H₂O resulting in the formation of hydroxyl radicals (${}^{ \cdot }{\text{OH}}$) and the ozonize anion (${\text{O}}_{3}^{\text{-}}$) (Eq. 15.7a), or a two-electron transfer resulting in the formation of the hydroperoxide anion (${\text{HO}}_{2}^{\text{-}}$) and oxygen (O₂) (Eq. 15.7b). Following this initial step, propagation reactions with ozone and the ozonize anion yield the hydroperoxyl radical (${\text{HO}}_{{2}}^{ \cdot }$) and superoxide radical (${\text{O}}_{{2}}^{{ \cdot {\text{-}}}}$) (Eqs. 15.7c–15.7i). This suite of strong radicals is similar to those generated in the CHP system (Eqs. 15.5a–15.5h).

$${\text{O}}_{{3}} {\text{ + OH}}^{\text{-}} \to {\text{O}}_{3}^{\text{-}} { + }{}^{ \cdot }{\text{OH}}$$

(15.7a)

$${\text{O}}_{{3}} + {\text{OH}}^{\text{-}} \to {\text{O}}_{{2}} {\text{ + HO}}_{{2}}^{\text{-}}$$

(15.7b)

$${\text{O}}_{{3}} + {\text{HO}}_{{2}}^{\text{-}} \to {\text{O}}_{3}^{\text{-}} {\text{ + HO}}_{{2}}^{ \cdot }$$

(15.7c)

$${\text{HO}}_{{2}}^{ \cdot } \leftrightarrow {\text{O}}_{{2}}^{{ \cdot {\text{-}}}} + {\text{H}}^{ + }$$

(15.7d)

$${\text{O}}_{{3}} + {\text{O}}_{{2}}^{{ \cdot {\text{-}}}} \to {\text{O}}_{3}^{\text{-}} + {\text{O}}_{{2}}$$

(15.7e)

$${\text{O}}_{3}^{\text{-}} + {\text{H}}_{{2}} {\text{O}} \to {}^{ \cdot }{\text{OH}} + {\text{OH}}^{\text{-}} + {\text{O}}_{{2}}$$

(15.7f)

$${\text{O}}_{3}^{\text{-}} + {}^{ \cdot }{\text{OH}} \to {\text{O}}_{{2}}^{{ \cdot {\text{-}}}} + {\text{HO}}_{{2}}^{ \cdot }$$

(15.7g)

$${\text{O}}_{3}^{\text{-}} + {}^{ \cdot }{\text{OH}} \to {\text{O}}_{{3}} {\text{ + OH}}^{\text{-}}$$

(15.7h)

$${}^{ \cdot }{\text{OH}} + {\text{O}}_{{3}} \to {\text{HO}}_{{2}}^{ \cdot } + {\text{O}}_{{2}}$$

(15.7i)

Ozone can react with target PHC compounds either through direction oxidation, or through the free radicals generated. Direct oxidation mechanisms include 1,3-dipolar cycloaddition of ozone to an alkene bond (a carbon–carbon double bond), and electrophilic attack on aromatic hydrocarbons (Smith and March 2007; Langlais et al. 1991).

Ozone has a long history of being used to treat PHCs associated with fuel releases. The use of IAS coupled with SVE has been used to remove mass from LNAPL source zones by extracting volatile compounds and stimulating aerobic biodegradation (Bouchard et al. 2018). As treatment progresses and most of the assessable volatile and semi-volatile compounds are depleted, the cumulative mass removal by an IAS/SVE system becomes asymptotic. At this point in the remedial effort, ozone can be used as an oxidizing gas to aggressively attack the remaining recalcitrant LNAPL mass. Investigations where ozone was injected into unsaturated columns packed with PHC impacted soils have yielded mass removal rates of 95% for crude oil (Dunn and Lunn 2002), and between 50 and 94% for diesel fuel (Yu et al. 2007; Jung et al. 2005). Since all these column systems had minimal water content the main degradation reactions were in the gas phase, presumably occurring at the soil/gas interfaces. The rates of reaction of the various PHC chains were noted to be similar; however, the C₁₀ fraction took longer to degrade since higher carbon aliphatic compounds were oxidized to shorter-chain compounds before being oxidized to lower molecular weight organic compounds and CO₂. The majority of the lower weight organic compounds are suspected to be easily biodegraded (e.g., alcohols, aldehydes, ketones, and carboxylic acids).

Ozone has been shown to effectively degrade BTEX compounds both in the laboratory and in the field (Bhuyan and Latin 2012; Garoma et al. 2008; Black 2001). The direct reaction of ozone with BTEX compounds is given by Eqs. (15.8a)–(15.8d):

$${\text{C}}_{{6}} {\text{H}}_{{6}} {\text{ + 11O}}_{{3}} \to {\text{6CO}}_{{2}} {\text{ + 3H}}_{{2}} {\text{O + 11O}}_{{2}}$$

(15.8a)

$${\text{C}}_{{7}} {\text{H}}_{{8}} {\text{ + 18O}}_{{3}} \to 7{\text{CO}}_{{2}} {\text{ + 4H}}_{{2}} {\text{O + 18O}}_{{2}}$$

(15.8b)

$${\text{C}}_{{8}} {\text{H}}_{{{10}}} {\text{ + 7O}}_{{3}} \to 8{\text{CO}}_{{2}} {\text{ + 5H}}_{{2}} {\text{O }}$$

(15.8c)

$${\text{C}}_{{8}} {\text{H}}_{{{10}}} {\text{ + 21O}}_{{3}} \to 8{\text{CO}}_{{2}} {\text{ + 5H}}_{{2}} {\text{O + 21O}}_{{2}}$$

(15.8d)

Since ozone generates the similar suite of radicals as in the CHP system, ozone is expected to rapidly oxidize other PHCs in addition to BTEX. Tengfei et al. (2016) used O₃ as a pre-treatment method for weathered crude oil impacted soil, and observed 50% TPH reduction. In parallel with this reduction was a 20 times increase in Chemical Oxygen Demand (COD), and a 4 times increase in the 5-day biochemical oxygen demand (BOD) indicating that some PHCs were converted to partly oxidized products that easily biodegraded.

Ozone is very reactive with PAHs, and thus is very effective for the removal of PAHs from soils. PAHs with a few benzene rings (e.g., naphthalene, fluorine, phenanthrene) can be oxidized to a greater extent than PAHs with a larger number of rings (e.g., pyrene, chrysene, benzo(a)pyrene) (e.g., Nam and Kukor 2000; Masten and Davies 1997). Evidence suggests that a large number of intermediates may be formed depending on the parent PAH, and reaction conditions (gas or aqueous phase, pH, etc.). As a result, a series of oxidation steps are required to completely mineralize the parent PAH as illustrated by Yao et al. (1998) for the oxidation of pyrene.

15.3.3 Permanganate

Permanganate (${\text{MnO}}_{{4}}^{\text{-}}$) has been used to treat both drinking water and wastewater (Eilbeck and Mattock 1987) for the removal of iron and manganese but has also been used for the treatment of trihalomethane precursors (Colthurst and Singer 1982) and phenolic wastes (Vella et al. 1990). ${\text{MnO}}_{{4}}^{\text{-}}$ is a selective oxidant since it will degrade some organic compounds (Waldemer and Tratnyek 2006). There are two forms of permanganate that are widely used: potassium permanganate (KMnO₄ a crystal solid) and sodium permanganate (NaMnO₄ a liquid). Both forms have the same reactivity; however, the solubility of KMnO₄ is 65 g/L at 20 °C, and if higher concentrations are required NaMnO₄ is used (solubility of 900 g/L at 20 °C).

The ${\text{MnO}}_{{4}}^{\text{-}}$ ion (E^o = 1.7 V) is a transition metal oxidant that reacts only through direct electron transfer rather than through any radical intermediates like in the CHP and ozone systems. As shown by the half-reactions in Eqs. (15.9a–15.9c), the number of electrons transferred depends on the system pH. In most natural groundwater systems, Eq. (15.9b) is dominant with ${\text{MnO}}_{{4}}^{\text{-}}$ undergoing a three-electron transfer.

$${\text{MnO}}_{{4}}^{\text{-}} {\text{ + 8H}}^{ + } {\text{ + 5e}}^{\text{-}} \to {\text{Mn}}^{{2 + }} + {\text{4H}}_{{2}} {\text{O}}\quad {\text{pH}} < 3.5$$

(15.9a)

$${\text{MnO}}_{{4}}^{\text{-}} {\text{ + 2H}}_{{2}} {\text{O + 3e}}^{\text{-}} \to {\text{MnO}}_{{2}} {\text{(s)}} + {\text{4OH}}^{\text{-}} \quad 3.5 < {\text{pH}} < 12$$

(15.9b)

$${\text{MnO}}_{{4}}^{\text{-}} {\text{ + e}}^{\text{-}} \to {\text{MnO}}_{4}^{2 \text{-} } \quad {\text{pH}} > 12$$

(15.9c)

An unfortunate by-product of Eq. (15.9b) is the production of solid manganese dioxide (MnO₂) (Pisarczyk 1995). MnO₂ forms at the point of reaction as colloid size amorphous solids which coalesce into larger particles. The generated MnO₂ solids have the potential to agglomerate in pore bodies and throats which can result in a reduction of hydraulic conductivity (Schroth et al. 2001), and non-aqueous phase liquid (NAPL)/water mass transfer (Li and Schwartz 2004; MacKinnon and Thomson 2002). Observations from laboratory experiments suggest that the accumulation of MnO₂ occurs in regions of high NAPL saturations where there is a greater NAPL/water interfacial area for mass transfer to occur. For example, MacKinnon and Thomson (2002) reported that after exposure of a perchloroethylene (PCE) pool to permanganate, the MnO₂ accumulation at the NAPL/water interface was so thick and consolidated that the MnO₂ had effectively cemented sand into MnO₂ “rocks” (Fig. 15.3). An equivalent LNAPL architecture to a PCE pool would be at the top of the smear zone where higher LNAPL saturations are expected to be present (see Fig. 15.1).

Fig. 15.3

MnO₂ crust formation at the interface of a NAPL pool in a 2-D model aquifer system used to study permanganate flushing to deplete a PCE DNAPL pool: a view into the 2-D model during the excavation process, and b several sections of the sand/silica flour interface removed from the model aquifer showing that the accumulation of MnO₂ had agglomerated, so that rock-like deposits of MnO₂ and sand were formed (from MacKinnon and Thomson 2002)

As a result of mineralization, CO₂ will be generated and depending on the system pH, there is the potential for CO₂ gas to evolve from the dissolved phase (Lee et al. 2003; MacKinnon and Thomson 2002; Schroth et al. 2001). In some situations, the CO_2(g) may fill pore bodies and decrease hydraulic conductivity and mass transfer, while in other situations volatile PHCs present in the system may partition into the CO_2(g) causing spontaneous expansion and upward mobilization of the gas phase (Mumford et al. 2008). This latter situation may result in uncontrolled releases of PHC gases.

As described by Waldemer and Tratnyek (2006), the oxidation of an aromatic compound is through attack on a benzylic C–H bond (a hydrocarbon bond on an alkyl side chain of the aromatic ring). Toluene, ethylbenzene, and the xylene isomers, all contain benzylic C–H bonds and react with permanganate at slow but appreciable rates. Since benzene has no benzylic C–H bonds, it is almost unreactive with permanganate. The stoichiometric reaction between ${\text{MnO}}_{{4}}^{\text{-}}$ and toluene, and between ${\text{MnO}}_{{4}}^{\text{-}}$ and ethylbenzene is given by

$${\text{C}}_{{7}} {\text{H}}_{{8}} {\text{ + 12MnO}}_{4}^{\text{-}} {\text{ + 12H}}^{ + } \, \to {\text{12MnO}}_{2} + 7{\text{CO}}_{{2}} {\text{ + 10H}}_{{2}} {\text{O}}$$

(15.10a)

$${\text{C}}_{{8}} {\text{H}}_{{{10}}} {\text{ + 14MnO}}_{4}^{\text{-}} {\text{ + 14H}}^{ + } \, \to {\text{14MnO}}_{2} + 8{\text{CO}}_{{2}} {\text{ + 12H}}_{{2}} {\text{O}}$$

(15.10b)

Table 15.3 provides the second-order reaction rate coefficients for the reaction between ${\text{MnO}}_{{4}}^{\text{-}}$ and BTEX compounds (Rudakov and Lobachev 2000). These reaction rate coefficients are substantially smaller than those listed in Table 15.2 for the oxidation of BTEX compounds by hydroxyl radicals.

Table 15.3 Second-order reaction rate coefficients for the reaction between permanganate and individual BTEX compounds (from Rudakov and Lobachev 2000)

Only some PAHs are susceptible to oxidation by ${\text{MnO}}_{{4}}^{\text{-}}$. Forsey et al. (2010) performed a comprehensive investigation into the reactivity and kinetics of ${\text{MnO}}_{{4}}^{\text{-}}$ toward PAHs. The oxidation of naphthalene, phenanthrene, chrysene, 1-methylnaphthalene, 2-methylnaphthalene, acenaphthene, fluorene, carbazole isopropylbenzene, and methylbenzene closely followed pseudo first-order reaction kinetics. The oxidation of pyrene was initially very rapid and then transitioned into a pseudo first-order rate at later time. Fluoranthene was only partially oxidized, and the oxidation of anthracene was too fast to be captured. Biphenyl, dibenzofuran, and tert-butylbenzene were non-reactive under the study conditions. The oxidation rate was shown to increase with the increasing number of polycyclic rings since less energy is required to overcome the aromatic character of a polycyclic ring than is required for benzene. Thus, the rate of oxidation increased in the series pyrene > phenanthrene > naphthalene.

15.3.4 Persulfate

Persulfate is used in many industrial processes and commercial products. For example, it is used as an initiator in the polymerization reaction during the manufacturing of latex and synthetic rubber, in the electronics industry to clean and microetch circuit boards, and in the pharmaceutical industry in the preparation of antibiotics. In general, the chemistry of persulfate is quite complex and not fully understood. The persulfate anion (${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$), also known as peroxydisulfate (PDS), is a strong oxidant (E^o = 2.1 V), and can react directly with some PHCs through a two-electron transfer process (Eq. 15.11a). Many important persulfate reactions occur when ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ reacts with other species (i.e., activated) to form a suite of free radicals (Lee et al. 2018; Block et al. 2004). Reactive species formed can include both oxidants and reductants. While activated persulfate leads to the potential degradation of a wider range of PHC compounds, additional ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ is consumed and often not efficiently. Various forms of persulfate are available with the most common being solid potassium persulfate (Na₂S₂O₈). At room temperature, ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ is quite stable at near-neutral pH (half-life of ~1800 days at pH of 7 and 20 °C), but under strong acidic conditions (pH < 0.3) ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ hydrolyzes to generate Caros’ acid (${\text{H}}_{{2}} {\text{SO}}_{5}$) (Eq. 15.11b). Under acidic conditions, Caros’ acid or peroxymonosulfuric acid oxidizes water into ${\text{H}}_{{2}} {\text{O}}_{{2}}$ (Eq. 15.11c) by transferring two-electrons, and the generated ${\text{H}}_{{2}} {\text{O}}_{{2}}$ is available to participate in the reactions given by Eqs. (15.5a)–(15.5h).

Persulfate can be activated by several different methods including the use of transition metals, heat, peroxide, or high alkaline conditions. Activation with dissolved transition metals (Fe²⁺, Fe³⁺, Ag⁺, Cu²⁺) is one of the most common methods and has parallels with Fenton’s Reagent (Eq. 15.5b) (e.g., House 1962; Anipsitakis and Dionysiou 2004). Ferrous iron (Fe²⁺) is the most frequently used transition metal and involves a one-electron transfer from Fe²⁺ to ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ resulting in the generation of the sulfate radical (${\text{SO}}_{{4}}^{{ \cdot {\text{-}}}}$) (Eq. 15.11d) (Kolthoff et al. 1951). The sulfate radical is a powerful, non-selective oxidant (E^o = 2.4 V). Heat activation of persulfate (Eq. 15.11e), which also yields ${\text{SO}}_{{4}}^{{ \cdot {\text{-}}}}$, has been investigated at temperatures between 30 and 60 °C (e.g., Huang et al. 2005). Persulfate activation by peroxide in a dual oxidant system (Sra et al. 2013a; Crimi and Taylor 2007) is speculated to be initiated when ${}^{ \cdot }{\text{OH}}$ reacts with ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ (Eq. 15.11f). Finally, alkaline activation of persulfate involves that addition of a concentrated base (e.g., NaOH) to raise the pH of the persulfate solution to 11 or 12. At this elevated pH, ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ undergoes base-catalyzed hydrolysis to generate ${\text{SO}}_{{4}}^{{ \cdot {\text{-}}}}$ and the superoxide radical (${\text{O}}_{{2}}^{{ \cdot {\text{-}}}}$) (Eq. 15.11g). The high concentration of ${\text{OH}}^{\text{-}}$ will react with ${\text{SO}}_{{4}}^{{ \cdot {\text{-}}}}$ to yield ${}^{ \cdot }{\text{OH}}$ (Eq. 15.11h). Finally, like H₂O₂, ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ can be catalyzed by certain minerals (oxides of Fe or Mn) associated with the aquifer materials. Of all the persulfate activation methods, alkaline activation is considered the most aggressive.

$${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }} {\text{ + 2e}}^{\text{-}} \to {\text{2SO}}_{{4}}^{{2 \text{-} }}$$

(15.11a)

$${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }} {\text{ + H}}_{{2}} {\text{O}} \to {\text{H}}_{{2}} {\text{SO}}_{{5}} {\text{ + SO}}_{4}^{2 \text{-} }$$

(15.11b)

$${\text{H}}_{{2}} {\text{SO}}_{{5}} + {\text{H}}_{{2}} {\text{O}} \to {\text{H}}_{{2}} {\text{SO}}_{{4}} + {\text{H}}_{{2}} {\text{O}}_{{2}}$$

(15.11c)

$${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }} {\text{ + Fe}}^{{2 + }} \to {\text{SO}}_{{4}}^{{{\text{-}} \cdot }} {\text{ + SO}}_{{4}}^{{2 \text{-} }} {\text{ + Fe}}^{{3 + }}$$

(15.11d)

$${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }} \mathop{\longrightarrow}\limits^{{{\text{hv}}}}{\text{ 2SO}}_{{4}}^{{{\text{-}} \cdot }}$$

(15.11e)

$${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }} + {}^{ \cdot }{\text{OH}} \to {\text{SO}}_{{4}}^{{{\text{-}} \cdot }} {\text{ + SO}}_{{4}}^{{2 \text{-} }} {\text{ + 1/2O}}_{{2}} {\text{ + H}}^{ + }$$

(15.11f)

$${\text{2S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }} {\text{ + 2H}}_{{2}} {\text{O }}\mathop{\longrightarrow}\limits^{{{\text{OH}}^{\text{-}} }}{\text{ SO}}_{{4}}^{{ \cdot {\text{-}}}} {\text{ + O}}_{{2}}^{{ \cdot {\text{-}}}} {\text{ + 3SO}}_{{4}}^{{2 \text{-} }} {\text{ + 4H}}^{ + }$$

(15.11g)

$${\text{SO}}_{{4}}^{{ \cdot {\text{-}}}} {\text{ + OH}}^{\text{-}} \to {\text{SO}}_{{4}}^{{2 \text{-} }} { + }{}^{ \cdot }{\text{OH}}$$

(15.11h)

Similar to the CHP system, there are non-productive or scavenging reactions involving carbonate, bicarbonate, chloride or sulfate ions present in most groundwater systems that compete with the target PHCs for the radicals generated in an activated ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ system [similar to Eqs. (15.6a–15.6d)].

The ability of ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ to degrade PHCs depends on the activation approach used, and the PHC mixture (e.g., crude oil vs. gasoline). In general, ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ can be used to treat BTEX and other aromatic PHCs. Huang et al. (2005) investigated the degradation of 59 compounds, including BTEX, using thermally activated ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ and concluded that compounds with carbon–carbon double bonds or with reactive functional groups bonded to benzene rings were readily degraded. Sra et al. (2013a) conducted a comprehensive investigation that focused on the treatability of dissolved gasoline compounds by persulfate. Specifically, unactivated and activated ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ (peroxide, chelated-ferrous, and alkaline) methods were used. The behavior of ten (10) gasoline compounds (benzene, toluene, ethylbenzene, o-xylene, m, p-xylene (meta- and para-xylene are reported together), 1,2,3-trimethylbenzene, 1,2,4-trimethylbenzene, 1,3,5-trimethylbenzene, and naphthalene) along with two carbon fractions (C₆–C₁₀, and C₁₀–C₁₆) and TPH were monitored. The use of unactivated persulfate resulted in almost complete oxidation of BTEX (> 99%), trimethylbenzenes (> 95%), and significant oxidation of naphthalene (∼ 70%). Observed first-order rate coefficients (k_obs) were enhanced by 2–15 times using either peroxide or chelated-iron activation methods. Alkaline activation at pH 11 or 13 yielded a k_obs that was ∼2 times higher than the unactivated case, except for BTE where the k_obs was reduced by 50% at pH of 13. The observed degradation trends for the lower carbon fraction (C₆–C₁₀) were consistent with BTEX behavior; however, the higher carbon fraction (C₁₀–C₁₆) did not show increased degradation by any of the persulfate activation methods explored. All trials demonstrated an initially fast destruction (<10 days) of the higher carbon fraction followed by a relatively slower oxidation rate or complete stalling suggesting the presence of recalcitrant PHCs. Overall, the oxidation of the higher carbon fraction was observed to be significant (60–85%) across most of the trials. The bulk gasoline stoichiometry for these experimental trials varied from 120 to 340 g-persulfate/g-TPH. The lower end of this stoichiometric range was for unactivated ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ and clearly highlights the non-productive consumption of ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ when activation methods are used. 

Using a simulated mixture of gasoline PHCs (BTEX, nitrobenzene, and naphthalene) and a variety of porous media, Crimi and Taylor (2007) concluded that the degradation of the selected PHCs varied by ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ activation method and porous medium used. The various activation methods resulted in different post-treatment oxidation gas chromatograph patterns suggesting that the organic by-products formed were dependent on the activation method used.

Sra et al. (2013b) conducted a pilot-scale experiment to investigate the fate and transport of an injected ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ solution, and the concomitant treatment of a gasoline residual source zone. The high-resolution data collected indicated a 40–80% reduction in mass discharge in the monitored PHC compounds following treatment.

The majority of studies that have investigated the use of ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ to treat diesel fuel have used fresh diesel fuel (e.g., Liang and Guo 2012; Do et al. 2010), and not weathered diesel fuel as expected to be encountered at most sites. Yen et al. (2011) used both soils spiked with fresh diesel fuel, and soil samples collected from a site impacted with diesel fuel to examine the ability of Fe activated ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ to remove mass (TPH). Similar to the observations reported by Sra et al. (2013a), for the oxidation of dissolved gasoline compounds by ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$, they observed an initial rapid decrease in the TPH soil concentration over the first 40 days of ~ 40% which was followed by an additional ~20% decrease over the next 120 days of treatment. Yen et al. (2011) hypothesized that slow rate of mass removal was due to the consumption of Fe²⁺, and the presence of recalcitrant PHCs.

Usman et al. (2012) investigated the capability of magnetite (Fe₃O₄) activated ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ to treat soil spiked with fresh crude oil and weathered crude oil. They reported that 73% of the weathered crude oil and 83% of the fresh crude oil was degraded in one week. Apul et al. (2016) screened the ability of several oxidants, including thermal activated ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ (45 °C), to degrade weathered crude oil present in two different soils. While the control (water only) removed ~40% of the TPH, the thermal activated ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ treatment removed about 70% of the TPH.

The degradation of naphthalene, the simplest PAH, in water and soil has been reported by Huang et al. (2005), Crimi and Taylor (2007), and Sra et al. (2013a) using a variety of ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ activation methods. Cuypers et al. (2000) used thermally activated ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ (70 °C) to successfully oxidize sorbed PAHs in soils and sediments. Zhao et al. (2013) examined the ability of thermal, citrate chelated Fe²⁺, alkaline, and H₂O₂ activated ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ to treat PAHs in soil samples collected from a coking plant. Thermal activation (40–60 °C) resulted in the highest removal of PAHs (90–99%), followed by Fe²⁺ activated (80–90%), H₂O₂ activated (65–80%), unactivated (~70%), and alkaline activated (55–70%). In contrast to the batch experimental systems used in most studies, Lemaire et al. (2013) reported on a series of column experiments performed where a H₂O₂ activated ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ solution was flushed into columns packed with homogenized soil collected from a former coking plant. The mass of PAHs removed (sum of the 16 US EPA regulated PAHs) was ~30%, presumably due to competition for the oxidants by the high carbonate and natural organic matter content soil.

15.4 Oxidant Interaction with Aquifer Materials

As discussed in Sects. 15.3.1 and 15.3.4, the presence of some minerals associated with aquifer materials will activate H₂O₂ and ${\text{S}}_{2} {\text{O}}_{8}^{2 \text{-}},$ and generate radical species at the mineral surface that may degrade some PHCs if present. The radicals are unlikely to diffuse very far (perhaps a few nanometers) from the surface but will, nevertheless, result in productive or beneficial oxidant consumption. Despite this benefit, at most sites the oxidant will contact uncontaminated porous media. Thus, an important consideration in the design of an ISCO system is the interaction of an oxidant with naturally occurring reductants and catalysts that are associated with uncontaminated aquifer material. If these interactions are significant, they will influence oxidant persistence and treatment efficiency following delivery into the TTZ.

Typically, the impact of the dissolved groundwater species is overshadowed by the aquifer solids. For example, Barcelona and Holm (1991) concluded that the oxidation capacity of dissolved phase species (Mn⁴⁺, Fe³⁺, and dissolved organic carbon) was insignificant compared to the measured oxidation capacity of the aquifer solids of samples collected from two sites. Inorganic species (e.g., minerals) containing iron (Fe), manganese (Mn), sulphur (S), and the natural organic matter (NOM) associated with the aquifer solids are usually of concern (Mumford et al. 2005). NOM may be both refractory and labile toward oxidation, and therefore, the oxidation of NOM is highly dependent on the reactivity of the various functional organic groups that comprise the NOM. The possibility of multiple inorganic species, as well as a range of NOM, creates an extremely heterogeneous environment in which reactions may occur (see Fig. 15.4).

Fig. 15.4

Adapted from Mumford et al. (2005)

Pore-scale conceptual model for natural oxidant interaction (NOI) showing the possibility of reaction with reduced aquifer solid species, reaction with dissolved NAPL species, and transport of un-reacted oxidant.

The result of the interaction between the selected oxidant and aquifer material leads to either a consumption of the oxidant by the aquifer solids, or an enhancement in the oxidant decomposition rate. When an oxidant is consumed, the reactive species associated with the aquifer solids are finite, and hence there exists a finite consumption or NOD (Natural Oxidant Demand). Once the maximum NOD is satisfied there is minimal additional reaction between the oxidant and the aquifer material. Conversely, an enhancement in the oxidant decomposition rate implies that there is infinite interaction capacity available.

To capture these behavioral differences and the associated underlying processes for all oxidant behavior, the term natural oxidant interaction (NOI) is used rather than natural oxidant demand (NOD). Regardless of the way in which NOI may manifest, it will decrease the mobility of the oxidant, the reaction rate with the target compound(s), and the mass of oxidant available thus creating an inefficient treatment system. Quantification of the NOI is a requirement for site-specific assessment and the design of cost-effective ISCO systems.

A general kinetic expression that can be used to capture the behavior of most oxidants in the presence of uncontaminated aquifer solids is given by Eq. (15.12):

$$\frac{{{\text{d[}}C_{{{\text{ox}}}} ]}}{{{\text{dt}}}} = \text{-} k_{{\text{o}}} {[}C_{{{\text{ox}}}} ] - \sum\limits_{{{\text{cat}}}} {k_{{{\text{cat}}}} [C_{{{\text{ox}}}} ]\,{[}C_{{{\text{cat}}}} ]} - \sum\limits_{{{\text{NOM}}}} {k_{{{\text{NOM}}}} [C_{{{\text{ox}}}} ]\,{[}C_{{{\text{NOM}}}} ]}$$

(15.12)

where C_ox, C_cat, and C_NOM is the concentration of the oxidant, catalyst and NOM respectively, and k_o, k_cat, and k_NOM is the reaction rate coefficient for a decomposition reaction, catalytic reaction(s), and NOM reaction(s) respectively. The first term on the right-hand side of Eq. (15.12) captures oxidant depletion due to auto or thermal decomposition processes, the second term represents oxidant reactions with catalysts (by definition the catalysts involved in these reactions are not consumed), and the third term represents reactions with the various forms of NOM.

Table 15.4 describes the observed NOI behavior of four commonly used oxidants in the presence of uncontaminated aquifer materials. The decomposition of H₂O₂ in porous media follows a pseudo first-order mass action law. Petri et al. (2011) presented first-order rate constants from 18 studies where the persistence of H₂O₂ was tracked. While a range of CHP systems and porous medium types were used, the median first-order decomposition rate coefficient was 0.2/h (half-life of 3.5 h). Ozone is highly reactive with NOM (e.g., Hsu and Masten 2001) as well as with reduced minerals (metal iron and manganese oxides) (Jung et al. 2004; Lin and Gurol 1998). In general, when exposed to O₃, the oxidizable fraction of the NOM is depleted giving rise to a finite NOD, but the iron and manganese oxides act like catalysts and hence have an infinite interaction capacity.

Table 15.4 Attributes of NOI behavior of the commonly used oxidants

In contrast to the behavior of H₂O₂ and O₃, ${\text{MnO}}_{{4}}^{\text{-}}$ concentration data collected from batch experiments indicate that, as time progresses, the concentration of ${\text{MnO}}_{{4}}^{\text{-}}$ decreases and asymptotically approaches a plateau (Xu and Thomson 2009). The decrease in ${\text{MnO}}_{{4}}^{\text{-}}$ concentration and the mass of aquifer material in the test reactor are used to estimate the ${\text{MnO}}_{{4}}^{\text{-}}$ NOD (mass of KMnO₄ consumed per mass of dried aquifer material in g/kg). NOD is analogous to the BOD which results from a test used to measure the amount of biodegradable organic content in water under aerobic conditions. Figure 15.5a shows a typical ${\text{MnO}}_{{4}}^{\text{-}}$ NOD temporal profile and the characteristic fast and slow consumption rates. A high degree of correlation was observed by Xu and Thomson (2009) between the maximum NOD (NOD_max) observed and the NOM content implying that organic carbon is the major reduced species contributing to permanganate consumption for many aquifer materials. Once the NOD_max has been satisfied, no additional ${\text{MnO}}_{{4}}^{\text{-}}$ will be consumed unless sufficient MnO_2(s) has been deposited on the aquifer materials giving rise to MnO_2(s) catalyzed decomposition (Xu and Thomson 2009; Stewart 1965). Figure 15.5b presents normalized breakthrough curves (BTCs) for a conservative tracer and ${\text{MnO}}_{{4}}^{\text{-}}$ from a column packed with uncontaminated aquifer material. The delayed arrival of ${\text{MnO}}_{{4}}^{\text{-}}$ is associated with the fast reaction rate, and the damped shape of the ${\text{MnO}}_{{4}}^{\text{-}}$ BTC reflects the slow reaction rate.

Fig. 15.5

a Typical permanganate NOD (g-KMnO₄/kg) profile collected from batch experiments (adapted from Xu and Thomson 2009). NOD_max is the maximum NOD observed. b Conservative tracer and permanganate normalized breakthrough curves from a column packed with uncontaminated aquifer material (adapted from Xu and Thomson 2006). The role that the fast and slow reactions play is indicated on both panels

The behavior of ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ is similar to H₂O₂ and O₃ with the decomposition of ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ following a first-order rate law (Sra et al. 2010) reflecting the interaction with catalysts and other reductants in the system. For example, Fig. 15.6 shows ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ degradation profiles in the presence of three aquifer materials. A second dose of ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ was added 125 days after the first dose. A first-order kinetic model was able to represent the ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ temporal profiles reasonably well. For all aquifer materials, k_obs decreased following the second ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ dose at Day 125 indicating that changes to the reactive capacity of the materials occurred during exposure to initial dose of ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$. The difference in k_obs values was attributed to the presence of a finite mass of oxidizable material that was initially consumed and gave rise to the accelerated ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ degradation observed at early time (Oliveira et al. 2016).

Fig. 15.6

Temporal persulfate concentration profiles following the exposure of three aquifer materials to a 1 g/L persulfate solution (adapted from Oliveira et al. 2016). A second persulfate dose (spike) was added at Day 125. The solid line is the best fit first-order kinetic model

To demonstrate the inherent differences between the persistence of H₂O₂, ${\text{MnO}}_{{4}}^{\text{-}}$, and ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ in the presence of aquifer materials, consider the following injection scenario: (1) an uncontaminated aquifer is subject to the injection of an oxidant (H₂O₂, ${\text{MnO}}_{{4}}^{\text{-}}$, or ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$) (Fig. 15.7a) in two sequential episodes spaced 30 days apart; (2) following injection the oxidant solution remains immobile as it reacts with the aquifer material; (3) the controlling in situ kinetic parameters are taken from bench-scale efforts performed on the same aquifer material (for H₂O₂ and ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ there was no finite mass of oxidizable material present); and (4) the injection concentration is adjusted so that the “oxidation strength” is identical for each oxidant. Figure 15.7b illustrates the oxidant concentration profiles over 60 days. The following observations are significant:

• H₂O₂ concentration is rapidly reduced.

• The H₂O₂ profiles following Injection 1 and Injection 2 are identical.

• The ${\text{MnO}}_{{4}}^{\text{-}}$ concentration profile following Injection 1 decreases quickly and then slows down as it approaches an asymptote.

• The decrease in the ${\text{MnO}}_{{4}}^{\text{-}}$ concentration profile following Injection 2 is much less than after Injection 1 reflecting the consumption of much of the fast-reacting NOM.

• The ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ concentration profile is identical following Injection 1 and Injection 2.

• Since the ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ reaction rate is substantially less than H₂O₂, an increase in ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ persistence occurs.

Fig. 15.7

Adapted from Siegrist et al. (2014)

a Injection of an oxidant into an uncontaminated aquifer. b Oxidant concentration profiles following two sequential pulse injection episodes into a synthetic aquifer for hydrogen peroxide (red), permanganate (purple), and persulfate (green).

15.5 Transport Considerations

For an ISCO system to be effective, contact is required between the delivered oxidant and the target PHC compounds. While a review of suitable oxidant delivery techniques is provided in Sect. 15.6.2, this section focuses on relevant oxidant transport considerations for solution-based oxidants (CHP, percarbonate, permanganate and persulfate). Ozone is delivered as a gas either into the unsaturated or saturated zone, and its transport behavior is controlled by multiphase flow theory (see Thomson and Johnson 2000).

Solution-based oxidants are frequently injected under elevated pressure and result in a rapid displacement of the pore water within the TTZ. During this injection phase, advective transport will dominate, and oxidant migration will be similar to a conservative tracer except it will react with both aquifer materials according to its natural oxidant interaction behavior (Table 15.4), and any oxidizable PHC compounds present. Groundwater velocities during this phase are expected to be substantially higher than under natural conditions. When catalysts or activators are co-injected with H₂O₂ or ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ to produce powerful radicals, these reactions occur immediately, and the radicals are short-lived. For example, the quasi steady-state concentration of ${}^{ \cdot }{\text{OH}}$ is of the order of 10^–12–10^–16 M, and the transport distance from the point of generation is only a few nanometers so they must be close to their target PHC to be effective. The injected oxidant solution will follow preferential pathways controlled by the presence of higher permeability or hydraulic conductivity (K) zones. Velocity variations within the TTZ will result in dispersion or apparent mixing of the oxidant solution. The generation of gases (e.g., O₂ in the CHP system) may alter flow pathways. At the interface of the dispersed oxidant solution and resident PHC laden groundwater, oxidation will occur based on the relative rates of reaction, and oxidant mass will be depleted.

After the injection phase is complete and elevated hydraulic pressures have dissipated to background levels, the injected oxidant solution will migrate with the natural groundwater flow. In order to maximize the rate of oxidation [see Eq. (15.1)] and provide sufficient oxidant mass to satisfy stoichiometric demands (e.g., see Sect. 15.6.1), and NOI requirements (Sect. 15.4), high oxidant concentrations are often used (50–100 g/L). These high concentrations result in a density difference between the oxidant solution and the surrounding groundwater, and give rise to a body force that will cause the oxidant solution to migrate downward. Schincariol and Schwartz (1990) show that significant instabilities in flow can occur even at density differences as low as 0.8 g/L. Density dependent transport of the oxidant solution plays an important role in its ultimate fate. As the oxidant solution migrates and mixes with groundwater by dispersion, the oxidant will react with target PHC compounds and be involved in NOI reactions. As a consequence, the oxidant will be consumed, and the spatial extent of migration affected. Persistent oxidants like permanganate and persulfate will perhaps be present for several weeks to months (if not consumed) while short-lived oxidants (H₂O₂) will be quickly depleted.

While the in situ destruction of target PHCs is one key attribute offered by ISCO, the second attribute is its capability to increase mass transfer from a LNAPL. Almost all of the reactions of interest that occur between an oxidant and a target PHC compound take place in the aqueous phase and not in the LNAPL. Therefore, LNAPL dissolution or mass transfer from the LNAPL to the aqueous phase is extremely important since it will, in part, control the degree of mass removal and hence treatment effectiveness. The rate at which PHC compounds dissolve from the LNAPL determines the dissolved phase plume concentrations and its longevity. The presence of a LNAPL results in dissolved phase PHC compounds that create a concentration gradient across a stagnant boundary layer between the LNAPL and the bulk aqueous solution (Fig. 15.8). The mass transfer of a PHC compound in the LNAPL into the aqueous phase is generally expressed by a macroscopic variation of the stagnate film model given by Schwarzenbach et al. 1993 (Eq. 15.13)

$$\frac{{{\text{d}}C_{w,i} }}{{{\text{d}}t}} = \text{-} k_{d} {(}C_{w,i}^{{{\text{eff}}}} - C_{w,i} )$$

(15.13)

Fig. 15.8

Schematic of dissolved phase concentration gradients near a LNAPL/aqueous phase interface with and without an oxidant present in close proximity. As the oxidant and reaction front migrate toward the LNAPL/aqueous phase interface, the concentration gradient steepens (blue arrow) and mass transfer is enhanced

where C_w,i is the aqueous phase concentration of the ith PHC compound in the LNAPL, $C_{w,i}^{{{\text{eff}}}}$ is the effective aqueous solubility estimated from modified Raoult’s law (Lee et al. 1992), and K_d is the lumped mass transfer or dissolution rate coefficient. In a porous medium, k_d cannot be determined from first principles, and hence various methods have been developed to estimate it from system parameters such as pore size distribution metrics, LNAPL saturation, molecular diffusion coefficient, and groundwater velocity (e.g., Powers et al. 1994). As shown by Eq. (15.13), the rate at which LNAPL mass is depleted is a product of k_d, and the concentration gradient between $C_{w,i}^{{{\text{eff}}}}$ and $C_{w,i}^{{}}$.

Oxidation reactions in the aqueous phase decrease the concentration of the PHC compounds in the bulk solution, and thus increase or steepen the concentration gradient (Fig. 15.8), which in turn will increase the overall rate of LNAPL mass removed from the system. Results from laboratory studies have shown that a 6–10 times increase in the mass transfer rate is possible to achieve when NAPL is exposed to an oxidant (Schnarr et al. 1998; MacKinnon and Thomson 2002). For enhanced mass transfer to be possible, the oxidant needs to be present near the LNAPL/aqueous phase interface.

In regions of the entrapped LNAPL source zone where preferential flow pathways exist, diffusive dominated transport is the principal mechanism for the migration of LNAPL compounds out of the non-advective zones and for the migration of oxidants into these zones. Persistent oxidants, those that are stable in the subsurface for an extended period of time, are particularly advantageous in these situations. When delivered into a TTZ by way of preferential flow pathways, these oxidants can diffuse from a preferential pathway toward more stagnant regions, driven by the oxidant concentration gradient as shown conceptually in Fig. 15.9. This is in the opposite direction as the gradient associated with the aqueous LNAPL compounds. This counter-diffusion or two-way diffusion process increases diffusive mass transfer by decreasing the distance between the LNAPL/aqueous phase interface and the area where the dissolved LNAPL compounds have been reduced as a result of oxidation.

Fig. 15.9

Illustration of a two-way diffusion process occurring between a preferential flow pathway and entrapped LNAPL mass. The oxidant must be available in the preferential flow pathway for sufficient time for diffusive transport to occur

Similarly, for chemical oxidants to react with PHC mass stored (dissolved and sorbed) in lower K zones, the oxidant must diffuse from preferential flow zones. Since diffusion is a slow transport process compared to advection, a persistent oxidant is required. For example, Cavanagh et al. (2014) investigated the effectiveness of flushing unactivated ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ in a higher K zone as a means to reduce BTEX mass flux from an adjacent lower K layer. After ~210 days, ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ had diffused 60 cm into the lower K layer with a concomitant reduction in BTEX mass flux between 95 and 99%.

15.6 Delivery and Mixing Techniques

The success of an ISCO system depends on the ability to deliver the selected oxidant into the subsurface so that the oxidant will contact the PHC compounds in the TTZ. For this contact to result in a favorable outcome (i.e., oxidation to non-toxic end products, or by-products that are susceptible to biodegradation), this contact needs to be made with a sufficient concentration of oxidant present and for a sufficient duration that will allow the reaction to proceed to completion. Many sites will require several oxidant delivery episodes to achieve the desired treatment outcome. The delivery approach will vary depending on the defined TTZ and remedial objectives. As described in Sect. 15.2, the TTZ may be the residual LNAPL in the unsaturated zone, the LNAPL mass in the saturated capillary fringe and the saturated groundwater zone (smear zone), or the dissolved phase plume. Regardless of the TTZ, the mass of oxidant delivered should be sufficient to degrade the target PHCs as well as satisfy the NOI requirements.

15.6.1 Oxidant Dosing and Volume

To estimate the target PHC mass requires that adequate spatial (vertical and horizontal) site characterization data are available, and thus a conceptual site model (CSM) that is sufficiently robust must exist. Specifically, geologic and hydrogeologic as well as the PHC distribution (LNAPL, dissolved, and sorbed) including physical and chemical composition must be established. At some sites the CSM may be insufficiently developed and additional site characterization data specific to support the design of an ISCO system must be gathered. This is extremely important at sites with a high degree of physical heterogeneities. CSM uncertainties must be acknowledged and taken into consideration.

The oxidant mass required per mass of PHC depends on the functionality of the oxidant selected and the catalyst or activator used. Published stoichiometric data, if available, can provide an initial estimate, but bench-scale treatability testing using several samples collected from within the TTZ is highly recommended. NOI behavior can also be quantified using bench-scale testing on non-impacted aquifer material samples (e.g., Sra et al. 2010; Xu and Thomson 2008, 2009). Scale-up from bench-scale results to field-scale can be refined and optimized based on pilot test data and supported by experience. Often bench-scale tests are not conducted at representative in situ oxidant to solids mass (M_ox/s) ratios, and thus extracted parameters need to be scaled appropriately. In addition, consideration should be given to the various rates that affect oxidant mass following delivery into a TTZ; these include:

• dissolution rate from the multicomponent LNAPL (if present) to the aqueous phase,

• PHC desorption rate from aquifer materials to the aqueous phase,

• reaction kinetics between the oxidant and dissolved PHCs,

• reaction kinetics between the oxidant and the aquifer materials bearing in mind the different ways that NOI can manifest,

• the rate of oxidant decomposition through non-radical-producing pathways (applicable to H₂O₂ and ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$), and

• the PV exchange rate due to groundwater flow which will affect the oxidant residence time.

As an example of an estimate for permanganate dosing, consider a 1 m³ block of porous medium with a mobile porosity of 30%, bulk density of 1800 kg/m³, toluene saturation of 1%, and NOI of 5 g-${\text{KMnO}}_{{4}}^{{}}$/kg. Using the stoichiometry provided by Eq. (15.10a), 15.5 g of ${\text{MnO}}_{{4}}^{\text{-}}$ or 20.6 g of ${\text{KMnO}}_{{4}}^{{}}$ is required for complete mineralization of each g of toluene (C₇H₈). Therefore, to mineralize the 2.6 kg toluene (density of 0.87 g/mL) in this porous medium block requires 53.5 kg of ${\text{KMnO}}_{{4}}^{{}}$, and to satisfy the NOI, and an additional 9 kg of ${\text{KMnO}}_{{4}}^{{}}$ is required for a total of 62.5 kg. To meet this oxidant dosing requirement, ~7 pore volumes (PVs) of solution with a concentration of 30 g-${\text{KMnO}}_{{4}}^{{}}$/L needs to be delivered. This estimate assumes that all the oxidant solution remains in the porous medium block following delivery, and ideal conditions exist including no other oxidant demands and NAPL/water mass transfer limitations. In addition, in this example it was assumed that the NOI value provided for ${\text{KMnO}}_{{4}}^{{}}$ was NOD_max and representative of the finite consumption by the aquifer materials. In the case where the NOI is infinite, then an approximate oxidant residence time needs to be determined and used to estimate the mass associated with NOI over this time frame.

This dosing example clearly illustrates that several delivery episodes are likely required to supply enough oxidant mass to a TTZ. Following a delivery episode, rebound of PHC concentrations in the aqueous phase will indicate the presence of remaining PHC mass in the TTZ. After several delivery episodes when the rate of PHC mass removal diminishes and performance monitoring data (e.g., soil cores) indicate that the majority of PHC mass has been depleted, consideration should be given to a transition from ISCO to bioremediation. Bioremediation is often more cost-effective after ISCO has substantially decreased the mass in the TTZ.

Ideally, the volume of oxidant solution delivered should be equal or larger than the TTZ pore volume (PV_TTZ) to ensure maximum coverage. The PV_TTZ can be estimated by multiplying the TTZ area, the sum of the thicknesses of the higher K layers within the TTZ (if appropriate), and the mobile porosity. The higher K regions should be identified within the CSM. The mobile porosity is the portion of the total porosity that participates in advective flow and can be estimated from tracer test data (see Payne et al. 2008). As a result of preferential flow pathways, non-uniform oxidant distribution will occur at most sites. Therefore, the delivered oxidant solution volume needs to be > 1 PV_TTZ to attain the desired oxidant distribution across the TTZ. At sites with large variations in K within the TTZ, strategic delivery approaches may be required. Given the uncertainty inherent in a CSM and the capability of the selected delivery method to meet expectations, the oxidant distribution should be monitored during delivery. For example, Stevenson et al. (2020) used the elevated electrical conductivity (EC) signature of an activated ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ solution to track its arrival and persistence in real time using a network of simple resistivity probes. This information can be used to refine and optimize the delivery strategy.

15.6.2 Delivery Methods

As shown schematically in Fig. 15.1, a TTZ for PHCs is typically located above the water table (unsaturated zone), and/or within the shallow saturated groundwater zone including the saturated capillary fringe. The delivery of solution-based oxidants in the unsaturated zone, either gravity-fed or injected under pressure, is associated with a number of challenges. First, solutions tend to migrate downward from their point of release due to gravity, which subsequently results in limited lateral spreading and a restricted radius of influence (ROI), and second, solutions will generally prefer to migrate along preferential paths. These difficulties lead to limited treatment and perhaps a loss of oxidant mass below a residual LNAPL TTZ in the unsaturated zone. As a result, the use of injection points (permanent or temporary) is typically restricted to the delivery of solution-based oxidants into a TTZ in the saturated zone.

As discussed in Sect. 15.3, H₂O₂ requires a catalyst, and ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ may require an activator (if needed). Since H₂O₂ and the catalyst react very fast, and radicals are short lived, the mixing of these two reactants together should ideally occur in the TTZ. There are two general approaches used: (1) the catalyst and H₂O₂ solutions are co-injected, and mixing takes place at the injection point, and (2) the catalyst solution is injected first and is followed by the H₂O₂ solution and the catalytic reaction takes place at the interface of the two solutions. Since ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ is more stable than H₂O₂, some of the species used to activate ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ (e.g., NaOH for high pH) are mixed with the ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ solution on ground surface just prior to injection.

15.6.2.1 Surface Application and Infiltration Trenches

For a residual LNAPL TTZ in the unsaturated zone, the use of ozone, the surface application of oxidants, or infiltration trenches are potential methods. Ozone is delivered as a gas either into the unsaturated or saturated zone using a distribution of sparge wells or points. A substantial volume of literature exists on the use of IAS to deliver air and other gases (e.g., see Johnson and Johnson 2012).

The surface application of oxidants may involve mixing a solid oxidant (${\text{MnO}}_{{4}}^{\text{-}}$ or ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$) into the top soil layer or irrigating an oxidant solution on the land surface above the TTZ. Following this initial application, infiltrating precipitation will dissolve the solids or mix with the applied oxidation solution and transport oxidant laden water downward through the unsaturated zone where it will contact the residual LNAPL. Clearly a surface application method requires the use of a persistent oxidant. A schematic of an infiltration trench or gallery to deliver an oxidant solution into a shallow LNAPL TTZ is shown in Fig. 15.10. In this approach, an oxidant solution is prepared in batch-mode in a surface facility, and then released into the system periodically. The solution is then distributed into a network of slotted pipes installed in high K material. Migration of the oxidant solution out of the infiltration trench is controlled initially by the pressure head in the slotted pipe network, but ultimately by gravity as it infiltrates through the unsaturated zone and into the TTZ.

Fig. 15.10

Schematic of an infiltration trench used to deliver an oxidant solution into a shallow LNAPL TTZ in the unsaturated zone

15.6.2.2 Injection Points

The most common method to deliver a solution-based oxidant is through the use of temporary or permanent injection points. Temporary injection points typically involve Direct Push Technology (DPT) where a small-diameter rod equipped with an injection point is pushed to the desired depth interval and then a specified volume of oxidant solution is injected (see Fig. 15.11a). The screened or slotted open section of a temporary injection point is typically 30–150 cm long based on the tooling used and vertical target depth interval desired.

Fig. 15.11

Oxidant distribution a following an injection at a target depth interval using DPT, b using the top-down mode of injection at four target depth intervals, and c using the bottom-up mode of injection at four target depth intervals

Depending on the mode of injection (top down or bottom up), the injection point is then advanced to the next depth interval and a specified volume of oxidant solution is again injected. This sequence is repeated for each target depth interval at that location. Figure 15.11b illustrates the top-down injection mode where the process starts at the shallowest depth interval and progresses down. The bottom-up mode of injection is quite similar to the top-down mode; however, in this case the injection point is initially advanced to the deepest depth injection interval and then sequentially pulled-up to the next shallowest depth interval (Fig. 15.11c). In the bottom-up mode, the borehole created by the rods (if it remains open) may act as a pathway for the downward migration of oxidant solution injected at shallower target depth intervals producing an oxidant distribution as shown in Fig. 15.11c. Following injection at all the depth intervals at a given location, and regardless of the injection mode, the injected oxidant solution will be subjected to density-driven advection if it has sufficient density contrast (see Fig. 15.12).

Fig. 15.12

Schematic illustrating the oxidant distribution following a initial top-down injection at four target depth intervals, and b, c the effects of density-driven transport resulting in the coalescence of the initial discrete depth intervals of oxidant

In contrast, a permanent injection point is an injection well (e.g., continuous slotted screen, large open area, coarse sand pack) with the capability to accept large flow rates. The length of an injection well screen is usually 300–460 cm. Permanent injection points are used in situations where multiple injection episodes are expected at the same location and depth interval, and remobilization expenses are prohibitive. Once installed, permanent injection points provide less flexibility to respond to monitoring data compared to temporary injection points since the injection of additional episodes would be at the same location and depth interval. Temporary injection points provide the opportunity to strategically target “hot spots” identified in the CSM, partially overcome geological controls (i.e., preferential flow pathways), and allow for maximum flexibility for additional injection episodes to focus on areas of rebound. In addition, fouling of the screened interval by reaction by-products is less likely to occur in temporary injection points compared to permanent injection points which are used repeatedly. Typically, the injected oxidant volume delivered at a temporary injection point is much smaller than the volume delivered into a permanent injection well.

Under typical injection pressures, an oxidant solution is incompressible and when injected into an aquifer the pore water, which is also incompressible, is displaced. As a result, the injected solution will migrate radially away from the injection point and into the surrounding aquifer until injection has stopped and the hydraulic pressure increase has dissipated. The transitory pressure field that develops as a result of the injection will result in a groundwater mound since in a shallow groundwater system this is the path of minimal hydraulic resistance. The three-dimensional shape of the oxidant solution as it exits the injection point will be controlled by the near-well K-field with the solution entering the higher K zones. Figure 15.13 illustrates the oxidant distribution from a permanent injection well into a homogeneous aquifer, and layered heterogeneous aquifer. While the oxidant distribution around the injection well is uniform for a homogeneous aquifer, the oxidant distribution around the injection well for the layered heterogeneous aquifer shows preferential migration along the higher K layers. These higher K layers provide the path of least resistance from the wellbore into the porous medium.

Fig. 15.13

a Oxidant distribution following injection into a homogeneous aquifer, and b layered heterogeneous aquifer from a permanent injection point. The theoretical radius of influence (ROI) is shown on panel (a) for the homogeneous aquifer

For a TTZ within a source zone, the injection wells are typically located on an off-set grid pattern with overlapping ROI as shown in Fig. 15.14. An ideal ROI can be estimated from Eq. (15.14):

$${\text{ROI}} = \sqrt {\frac{{V_{{{\text{inj}}}} }}{{\pi h\theta_{{\text{m}}} }}}$$

(15.14)

Fig. 15.14

Schematic layout of 25 injection points across a source zone TTZ. The solid symbol (●) represents the location of an injection point, and the circle around each location represents the expected ROI for each injection point. The ROI at each injection point overlaps neighboring ROIs to ensure adequate distribution of the injected oxidant solution

where V_inj is the injected volume of oxidant solution, h is the injection zone thickness (normally length of the well screen), and θ_m is the mobile porosity. Again, the mobile porosity is the portion of the total porosity that participates in advective flow (Payne et al. 2008). The overlapping ROIs account for non-ideal conditions and to prevent dead zones between each injection point. A review of monitoring data collected during the injection phase is essential to ensure that the spatial distribution of the oxidant can be understood. Normally the sequence of injections starts at the perimeter locations and advances inward to avoid displacing PHC laden groundwater outside the TTZ.

Delivery of an oxidant solution into a dissolved phase TTZ has additional challenges compared to a source zone TTZ. In a residual or entrapped source zone TTZ, the PHC mass is essentially immobile, while in a dissolved phase plume the PHC mass is a combination of mass in the aqueous and sorbed phases (depending on partitioning). In cases when the sorbed mass associated with the dissolved plume is minimal, the large volumes of oxidant solution injected will simply displace the target PHC laden groundwater since mixing is limited. Injection strategies that start at the edge of the plume and work inward will contain the PHC laden groundwater but will not enhance mixing. Following injection, and only if the oxidant is persistent, an oxidant slug will be transported by the ambient groundwater flow field and mixing along the edges will occur due to dispersion. Transverse dispersion (lateral or vertical) is weak compared to longitudinal dispersion (Gelhar et al. 1992) resulting in the majority of mixing occurring along the principal flow direction.

One potential approach to improve oxidant distribution in a dissolved phase plume is to use a cross-injection system (CIS) (Shayan et al. 2017; Gierczak et al. 2007; Devlin and Barker 1996). A CIS involves a series of injection-extraction wells used to deliver an oxidant solution perpendicular to the ambient flow direction in pulses (Fig. 15.15). Once the oxidant pulse between an injection well (IW) and extraction wells (EWs) has been established, the pumping system is stopped. The oxidant slug is then transported by the ambient flow field where dispersion will lead to spreading and mixing of the oxidant with PHC laden groundwater. In the example illustrated in Fig. 15.15, CIS would need to operate until such time at the source zone was treated.

Fig. 15.15

Example of a CIS used to deliver an oxidant into a dissolved phase plume showing: a the location of CIS wells in the plume, and b the IW and EW, the initial oxidant pulse between the IW and EWs, and the spreading and mixing of the oxidant pulse

To sustain a continual supply of oxidant solution into a TTZ, a recirculation system can be used as shown in Fig. 15.16 (see also Thomson et al. 2007; Lowe et al. 2002). This system is comprised of a number of injection and extraction wells, and a surface treatment system to amend the oxidant concentration, and remove problematic ions and by-products (e.g., Fe, MnO₂) that may foul injection wells. Some jurisdictions may have regulations that do not allow the injection of contaminated water, and hence any residual PHC compounds in the extracted groundwater would also need to be removed prior to re-injection. The continual supply of an oxidant solution ensures that reaction rates remain elevated, the rate of diffusion maximized, and adequate oxidant mass is being delivered to satisfy dosing requirements.

Fig. 15.16

Schematic cross-section of a recirculation system to maintain a continual supply of oxidant into a LNAPL TTZ. The treatment system amends the oxidant concentration and removes problematic compounds (PHCs and ions)

15.6.2.3 Other Methods

Electrokinetic Delivery

The inability to inject liquid oxidants into lower K TTZs significantly impacts the selection of ISCO as a preferred treatment technology. To overcome this limitation, electrokinetic (EK) enhanced delivery of chemical oxidants has been proposed (termed EK-ISCO) (Xu et al. 2020; Chowdhury et al. 2017a; Pham and Sillanpää 2015; Fan et al. 2014; Reddy 2013; Reynolds et al. 2008; Roach and Reddy 2006). The EK delivery principle involves the application of a low-intensity DC voltage to an array of electrode wells installed in a low-K TTZ. Negatively charged ions (e.g., ${\text{MnO}}_{{4}}^{\text{-}}$, ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$) supplied at the cathodes (negative) will migrate to the anodes (positive) and be uniformly distributed across the porous medium between electrodes regardless of the K heterogeneity (Wu et al. 2012). EK transport is a combination of electromigration (movement of ions in response to the applied electric field) and electroosmosis (movement of pore fluid in response to the applied electrical field from the anode to the cathode) (Acar and Alshawabkeh 1993). The DC voltage gradient applied across the electrode field is the driving force for EK transport. A pH buffering system is required at each electrode well since the electrolysis of water will produce H⁺ at the anode, and OH⁻ at the cathode. Cox et al. (2021) reported on a field demonstration designed to assess the ability of EK to delivery ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ into a clay unit. Chowdhury et al. (2017b) described on a proof-of-concept laboratory experiment that extends EK-ISCO to include a thermal ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ activation step. In this approach, ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ is initially delivered by EK into a TTZ, and then the DC voltage is switched to AC, and the resistance of the aquifer material causes heating of the system. This activation step is similar to electric resistive heating (ERH, Beyke and Fleming 2005); however, the target temperature is lower than 90–100 °C.

Slow-Release Oxidant Materials

At some sites, it might be advantageous to control the rate of oxidant delivery. To meet this need a variety of slow-release oxidant materials have been developed to provide low-cost, passive, and long-term treatment. Since their initial introduction (Lee and Schwartz 2007; Ross et al. 2005; Kang et al. 2004) more than 20 laboratory studies have been published using waxes (e.g., paraffin), gels (e.g., chitosan), or a mixture of silica sand, cement, and bentonite as binding materials (O’Connor et al. 2018). Encapsulated in the binding material are ${\text{MnO}}_{{4}}^{\text{-}}$ or ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ solids. The rate of oxidant release (dissolution) depends on the binding material and the morphology (i.e., candles, pellets, or gels). Slow-release materials shaped like candles can be placed in a series of wells strategically located within a TTZ, or as the reactive media in a permeable reactive barrier (PRB) used to cut-off a dissolved phase plume. In either case, oxidant delivery is controlled primary by diffusion out of the well casing and into the surrounding groundwater. As the concentrated oxidant plume migrates downgradient it will slowly mix and react with the surrounding PHC laden groundwater. A major limitation of this type of slow-release delivery system is its narrow zone of influence associated with each release well (O’Connor et al. 2018). To expand the zone of influence, Christenson et al. (2012) used an airlift pump system to generate a continuous circulation pattern around each well thereby increasing the spreading of the oxidant as it dissolves from the slow-release material. Reece et al. (2020) present data from three field sites where this oxidant delivery system has been implemented. Slow-release gel materials can be injected directly into a TTZ where they can release low concentrations of oxidant for extended periods of time (e.g., Yang et al. 2016; Lee and Gupta 2014).

15.6.3 Mixing

Following the injection of an oxidant solution into a TTZ, the desired chemical reactions are initiated by transport processes that bring the oxidant and aqueous PHC compound together (Kitanidis and McCarty 2012). This contact between the oxidant and an aqueous PHC compound is the desired outcome of an effective delivery system; however, it can be one of the most difficult to achieve. While spreading increases the zone of contact between the oxidant and aqueous PHC compound, physical heterogeneities create preferential flow pathways at a range of spatial scales that lead to uneven spatial distribution of the injected oxidant.

Ineffective oxidant delivery can lead to partial removal of the PHC mass. The remaining PHC mass remains as immobilized, disconnected pockets that are trapped in diffusion-controlled zones (Conrad et al. 1992) or in the lower K zones where contact with the injected oxidant is limited. In these regions, molecular diffusion is the primary transport mechanism responsible for bringing the oxidant and PHC compound together (Gale et al. 2015; Jose and Cirpka 2004); however, diffusion is a slow process (see Fig. 15.9).

As noted in Sect. 15.6.2, in conventional recirculation systems, the extracted groundwater is amended before being re-injected. Conventional recirculation systems typically use steady and uniform flows; however, a time-dependent flow field is required to promote enhanced mixing (Ottino 1989; Aref 1984). Chaotic advection has the potential to improve oxidant distribution and promote enhanced mixing (Aref et al. 2017; Trefry et al. 2012). Chaotic advection can be defined as the creation of “small-scale structures” caused by the repeated stretching and folding of fluid elements in a laminar flow regime (Aref et al. 2017; Ottino 1989). This rapid stretching and folding creates highly complex particle trajectories which can improve the spatial distribution of oxidants in porous media and increase the zone of contact where reactions can occur (Mays and Neupauer 2012; Ottino 1989). Cho et al. (2019) employed a Rotated Potential Mixing (RPM) flow configuration in a proof-of-concept field investigation and used multiple lines of evidence to demonstrate that enhanced mixing using chaotic advection is possible. This is the first reported attempt to investigate aspects of chaotic advection at the field scale. The findings from this proof-of-concept study represent a critical step in the development of a novel method that can be used to improve the effectiveness of ISCO.

15.7 Options for Integrated Treatment

There is a myriad of in situ treatment technologies available for deployment by remediation scientists and engineers. The diversity of these technologies is a testament that no single technology is ideal for all sites as each has its own set of strengths and weaknesses making it suitable for a given hydrogeologic setting. To minimize clean-up costs and time and maximize remedial efficacy, an integration or combination of technologies at sites is desired. ISCO can be combined or integrated with other treatment technologies, either concurrently or sequentially. The concurrent use of a technologies to treat the same TTZ requires a clear understanding of technical compatibility and expected synergies, while the sequentially use of technologies (i.e., treatment train) requires a prescription of key transition points.

It is well established that oxidation of some PHC compounds generates bioavailable substrates that lead to enhanced rates of biodegradation suggesting that ISCO will enhance subsequent microbial activity (Tengfei et al. 2016; Sahl and Munakata-Marr 2006). In general, microbial communities are inhibited during exposure to oxidants, but both the microbial concentration and diversity increase following oxidant exposure (e.g., Kim et al. 2021; Richardson et al. 2011; Tsitonaki et al. 2008). Bartlett et al. (2019) reported that sulfate-reducing bacteria and the associated microbial community rebounded to pre-exposure levels following exposure to unactivated or alkaline activated ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$.

Considering these impacts on the microbial community, the sequential use of either monitoring natural attenuation (MNA) or enhanced in situ bioremediation (EISB) following ISCO is warranted. As shown in Eqs. (15.11a–15.11h), the decomposition of ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ from reactions with PHC compounds or aquifer materials will generate excess amounts of sulfate (${\text{SO}}_{{4}}^{{2 \text{-} }}$). This excess ${\text{SO}}_{{4}}^{{2 \text{-} }}$ can serve as a terminal electron acceptor and enhance subsequent microbial sulfate reduction processes. Thus, following a ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ treatment episode the generated ${\text{SO}}_{{4}}^{{2 \text{-} }}$ plume will migrate downgradient and enhance microbial sulfate reduction. Understanding the interplay between ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ oxidation and subsequent enhanced microbial sulfate reduction is necessary to optimize the performance of this combined PHC treatment system (see Shayan et al. 2017). Similarly, peroxide, percarbonate, or ozone all generate a short-term footprint of O₂, a terminal electron acceptor for aerobic biodegradation. While the concentration of these oxidation end-products may be quickly depleted the delivery of additional electron acceptors may be required to sustain acceptable rates of biodegradation.

Surfactants and cosolvents can enhance the dissolution process, and will increase the concentration of PHC compounds in the aqueous phase where oxidation takes place (Wang et al. 2013; Li 2004). While a surfactant or cosolvent can be delivered prior to the oxidant, contact between the oxidant and the region of increased concentration of PHC compounds will suffer the same mixing limitations as discussed in Sect. 15.6.3. An alternate approach is to co-inject the surfactant or cosolvent along with the oxidant; however, both surfactants and cosolvents are organic compounds and are susceptible to oxidation. The compatibility of the surfactant or cosolvent with the selected oxidant is critical since this interaction will affect treatment effectiveness (see Dugan 2008, and Zhai et al. 2006).

Finally, sub-boiling thermal technologies can be used to thermally activate ${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$ following delivery. Increasing the temperature in the TTZ will also enhance mass transfer (e.g., Imhoff et al. 1997), and increase the rate of reaction between the oxidant and the PHCs as described by the Arrhenius law (Levenspiel 1999). Sustainable methods (e.g., solar or geothermal heating) that can strategically increase subsurface temperatures to the sub-boiling range are currently available (Horst et al. 2018).

15.8 Closure

This chapter has presented a number of key aspects of ISCO that are relevant to application at sites contaminated with PHCs. The following are important takeaway messages:

• A robust CSM is required. This CSM must include geologic and hydrogeologic data, and an understanding of the PHC mass distribution in each phase (LNAPL, dissolved, and sorbed). If the CSM is insufficiently developed, then additional site characterization data specific to support the design of an ISCO system must be gathered. The degree of physical heterogeneity (spatial variation in K) is critical. CSM uncertainties must be acknowledged and taken into consideration.

• The TTZ and ISCO treatment objectives must be clearly defined. A distinction between residual LNAPL in the unsaturated zone, LNAPL mass in the saturated capillary fringe and the saturated groundwater zone (smear zone), and the dissolved phase plume is required.

• ISCO should be considered part of integrated remediation strategy and not in isolation. Attempts to remove mobile LNAPL from the TTZ should be exhausted before ISCO is implemented. Once mass of PHCs oxidized becomes limited, transition to biodegradation is recommended.

• The common oxidants used are CHP, percarbonate (Na₂CO₃·1.5H₂O₂), ozone (O₃), permanganate (${\text{MnO}}_{{4}}^{\text{-}}$), and persulfate (${\text{S}}_{{2}} {\text{O}}_{{8}}^{{2 \text{-} }}$). O₃ is delivered as a gas, and the other four are typically delivered as concentrated solutions.

• A basic understanding of the chemistry and main reactions for each oxidant is required. CHP, percarbonate, ozone, and persulfate can generate a suite of highly reactive radical species.

• Most of the important PHC degradation reactions occur in the aqueous phase.

• In general, the commonly used oxidants can degrade target PHC compounds in the aqueous phase at acceptable reaction rates; however, some longer chain PHC compounds are recalcitrant.

• Permanganate is unable to oxidize benzene, a common risk driver at most sites.

• Problematic reaction by-products from oxidants include gases and/or solid precipitates that may decrease aquifer K and mass transfer from the LNAPL to the aqueous phase.

• Volatile PHCs present in the system may partition into the gases causing spontaneous expansion and upward mobilization of the gas phase leading to uncontrolled releases of PHC gases.

• Reaction of radicals with common groundwater ions (carbonates, chloride, and sulfate) will affect treatment efficiency.

• Reactions with uncontaminated aquifer materials result in either a finite consumption of the oxidant or an enhancement in the oxidant decomposition rate. The interaction of an oxidant with aquifer materials will decrease the mobility of the oxidant, the reaction rate with the target PHCs, and the mass of oxidant available thus creating an inefficient treatment system.

• Stoichiometric mass requirements can be substantial and must be appreciated. Activation methods increase the stoichiometric mass requirements.

• Sufficient oxidant mass needs to be delivered into a TTZ to meet the productive and non-productive demands.

• Preferential flow pathways will control the oxidant distribution in a TTZ.

• A persistent oxidant is needed to allow the oxidant to either diffuse away from preferential pathways or diffuse into lower K regions.

• A variety of oxidant delivery methods are available with temporary injection points being the most flexible to surgically deliver an oxidant to hot spots identified in the CSM, and respond to monitoring data.

• Ineffective oxidant delivery will lead to incomplete treatment of the TTZ since mixing is limited. Novel mixing methods based on time-dependent flows offer a technique that potentially overcomes this limitation.

• Opportunities exist to combine or integrate ISCO with other treatment technologies.