Recent Advances in 1,4-Dioxane Removal Technologies for Water and Wastewater Treatment

Abstract

1,4-Dioxane is a contaminant of emerging concern and a probable human carcinogen that has been widely detected in aqueous environments. However, the removal of 1,4-dioxane by conventional water and wastewater treatment plants had proven to be ineffective due to its unique physicochemical properties. The development of innovative technologies for both in-situ and ex-situ treatment of 1,4-dioxane to meet increasingly strict standards is in urgent need. This review summarizes the current available physicochemical and biological treatment technologies for the removal of 1,4-dioxane from both water and wastewater and the strategies that may potentially fulfill the stringent 1,4-dioxane standard were discussed. Advanced oxidation processes (AOPs), such as ultraviolet radiation coupled with H₂O₂ (8–10 mg L⁻¹), had shown efficient 1,4-dioxane destruction and had already been applied for both water and wastewater treatment processes. On the other hand, more than 30 pure microbial strains and microbial communities that can metabolically or metabolically degrade 1,4-dioxane were reported. Biodegradation has been proven to be a feasible and cost-effective approach for 1,4-dioxane remediation. Suspended growth bioreactor, immobilized cell bioreactor, and biofiltration systems were the most commonly used biological approaches to remove 1,4-dioxane from contaminated water. Though 1,4-dioxane easily desorbs after the adsorption by materials such as granular activated carbon (GAC) and zeolite, temporary 1,4-dioxane removal by adsorption followed by 1,4-dioxane biodegradation in the bioaugmented adsorption media may be a feasible strategy treating 1,4-dioxane contaminated water. Overall, the treatment chain that combines physical-chemical processes and biodegradation has a great potential for synergistic removal of 1,4-dioxane at lower operating costs.

1. Introduction

1,4-Dioxane is a contaminant of emerging concern that has been widely detected in both groundwater and surface water in the U.S. and worldwide. Historically, the majority (~90%) of 1,4-dioxane was used as a solvent stabilizer for chlorinated solvents, such as trichloroethylene (TCE) and 1,1,1-trichloroethane (1,1,1-TCA). Though TCA had been phased out under the Montreal Protocol and the application of 1,4-dioxane as solvent stabilizer was terminated [1,2], 1,4-dioxane remained as a chemical with high production volume, exceeding 1 million pounds per year [2,3]. It is currently used as a solvent for paints, dyes, greases, antifreeze, and aircraft deicing fluids, is used as a purifying agent in the manufacture of pharmaceuticals, and is a byproduct in the manufacture of polyethylene terephthalate (PET) plastic [2,4,5]. 1,4-Dioxane is also an impurity during the production of ethoxylated surfactants and can be detected in personal care and cleaning products such as deodorants, shampoos, and cosmetics [2,6].

The unique physiochemical properties of 1,4-dioxane resulted in its high mobility and persistence in the aqueous environment (Table 1). Its low Henry’s law constant (4.80 × 10⁻⁶ atm-m³ mol⁻¹ at 25 °C) and the complete miscibility in water result in low volatility in aqueous solutions and high mobility in water. The removal of 1,4-dioxane from water by soil or suspended solid via adsorption is limited based on the low octanol-water partition coefficient (log K_ow = −0.27) and low organic carbon partition coefficient (log K_oc = 1.23). Due to its high affinity to water and the ineffective removal by conventional wastewater treatment processes, 1,4-dioxane was widely detected in surface water and groundwater caused by accidental spillage, or improper disposal of industrial waste [7]. Data from the Unregulated Contaminated Monitoring Rule 3 (UCMR3) revealed 1,4-dioxane was detected in 21% of U.S. public water systems that utilize surface water or groundwater as the water source. The presence of 1,4-dioxane in the water system has raised global concerns due to the adverse health effect of the exposure. It is classified as a Class B2 (probable) human carcinogen by the International Agency for Research on Cancer (IARC) and the United States Environmental Protection Agency (USEPA) [8,9]. Acute exposure to 1,4-dioxane causes irritation of the eyes, nose, and throat, and exposure to elevated levels of 1,4-dioxane may cause kidney and liver damage in humans [4]. Chronic toxicity studies on rats had revealed general toxicity and increased nasal cavity, lever, kidney, and lung lesions after both long- and short-term chronic exposures [10,11].

The World Health Organization (WHO), as well as federal agencies of Japan, Korea, and Canada, suggested a guideline of 1,4-dioxane concentration (50 µg L⁻¹) in drinking water [12,13,14,15]. No federal maximum contamination level (MCL) had yet been set up for 1,4-dioxane in the U.S. USEPA set up a Health advisory level of 0.35 µg L⁻¹ for 1,4-dioxane in drinking water based on the cancer risk of 10⁻⁶ [9] and New York State adopted the first-in-the-nation MCL of 1,4-dioxane at 1 µg L⁻¹ in 2020 [16]. Several states had established their own drinking water criteria for 1,4-dioxane, with the concentration ranging from 0.3–7.2 µg L⁻¹ [17].

Table 1. Physical properties of 1,4-dioxane *.

Property	Value	Structure
Molecular weight	88.11 g mol−1
Density (25 °C)	1.033 g cm−3
Boiling point	101.1 °C
Water solubility	Miscible
Henry’s law constant at 25 °C	4.80 × 10−6 atm-m3 mol−1
Octanol-water partition coefficient (log Kow)	−0.27
Organic carbon partition coefficient (log Koc)	1.23

Note: * Data acquired from [18,19].

Table 1. Physical properties of 1,4-dioxane *.

Property	Value	Structure
Molecular weight	88.11 g mol−1
Density (25 °C)	1.033 g cm−3
Boiling point	101.1 °C
Water solubility	Miscible
Henry’s law constant at 25 °C	4.80 × 10−6 atm-m3 mol−1
Octanol-water partition coefficient (log Kow)	−0.27
Organic carbon partition coefficient (log Koc)	1.23

Note: * Data acquired from [18,19].

The overall goal of this work is to review current available 1,4-dioxane removal technologies based on a variety of removal mechanisms. The 1,4-dioxane removal performance by physical and chemical approaches, such as adsorption and advanced oxidation processes (AOPs) were discussed in depth. Furthermore, a comprehensive review of the biological 1,4-dioxane removal technologies was conducted with a focus on biodegradation pathways, degradation kinetics, and environmental impacts on degradation performance. Different bioreactor configurations were reviewed and discussed at various test scales. Finally, the challenges and opportunities of adopting 1,4-dioxane removal technologies for water and wastewater treatment were proposed and discussed.

2. Material and Methods

A systematic literature search of the peer-reviewed publications in the Web of Science was conducted to evaluate 1,4-dioxane removal techniques. The search included the keywords (1,4-dioxane) AND (adsorption OR advanced oxidation OR biodegradation OR bioreactor OR filter OR biofiltration OR chlorinated solvents OR natural attenuation OR metabolic OR co-metabolic). The literature search was limited to peer-reviewed publications written in English between 1990 to 2023. A total of 261,704 results can be searched from the database using the above keywords, and after a full-text review, 117 references passed our criteria. Study inclusion criteria including the scope and data availability were applied to each publication. To be considered within scope, the article needs to have the following information: for 1,4-dioxane removing techniques, information on 1,4-dioxane removal efficiencies, 1,4-dioxane removing kinetics, system configurations, and operating conditions were required for articles to pass the scope. For 1,4-dioxane degrading microorganisms, the 1,4-dioxane biodegradation kinetics and microbial growth kinetics, functional enzymes, and biodegradation pathways were required for articles to pass the scope. Publications reporting only presence/absence data were excluded. For each publication, the Monod kinetic coefficients of the 1,4-dioxane degrading microorganisms were determined by fitting the Monod equation:

$$\frac{dS}{dt}=-\frac{q_{max}XS}{K_s+S}$$

(1)

where S is the 1,4-dioxane concentration (mg L⁻¹), X is the biomass concentration measured as total protein (mg L⁻¹), t was time (h), q_max is the maximum 1,4-dioxane degradation rate (mg dioxane hr⁻¹ mg protein⁻¹), and K_s is the 1,4-dioxane half-velocity constant (mg L⁻¹). The removal efficiency of different systems was calculated using Equation (2) when studies reported the 1,4-dioxane removal efficiency:

$$Removal efficiency\left(\%\right)=\left[(\frac{C_{in}{-C}_{out}}{C_{in}})\times 100\right]$$

(2)

where C_in and C_out represent the influent and effluent 1,4-dioxane concentrations, respectively.

3. Physical-Chemical Processes for 1,4-Dioxane Removal

3.1. 1,4-Dioxane Removal by Adsorption

The removal of 1,4-dioxane by conventional water treatment processes, such as coagulation, flocculation, and air stripping, is generally considered ineffective due to the unique physicochemical properties of 1,4-dioxane [20,21]. Adsorption had been widely used for the removal of contaminants (e.g., pesticides, pharmaceuticals, and heavy metals) in aqueous environments [22,23]. Adsorption of 1,4-dioxane was considered ineffective due to the low K_ow and high water solubility. On the other hand, batch-scale adsorption kinetics and isotherm tests were conducted with a few conventional and modified adsorptive materials to evaluate their 1,4-dioxane removal potential (Table 2). Among the tested materials, Norit 1240^TM granular activated carbon (GAC) showed the maximum adsorption capacity (q_max) of 59.56 mg-1,4-dioxane g-GAC⁻¹ [24]. However, up to 50% 1,4-dioxane was easily desorbed after 5 sequential washing steps with fresh Ammonia Mineral Salt (AMS) medium. The adsorption ability of eight types of activated carbon was evaluated based on the 1,4-dioxane adsorption capacity at an equilibrium of 50 µg L⁻¹ (q₅₀). The q₅₀ of the ACs ranged from 4.3 to 410 µg g-AC⁻¹, with the highest adsorption capacity by the AC prepared from sawdust [25].

Rapid and effective 1,4-dioxane adsorption was also reported with zeolite-based materials [26,27,28]. The adsorption capacity of 1,4-dioxane on ZSM-5 zeolite with different Si/Al ratios ranged from 22.44 to 107.36 mg-1,4-dioxane g-zeolite⁻¹. The 1,4-dioxane adsorption capacity of ZSM-5 was reported to be linear with the surface area and the micropore volumes (Table 2). Higher Si content of zeolite also increased the hydrophobic siloxane linkage and thereby increased the 1,4-dioxane adsorption capacity [26,27]. Titanium silicate (TS) is a synthesized zeolite material with strong hydrophobic nature and selectivity [28]. 1,4-Dioxane was adsorbed by TS through the combination of hydrophobic interactions and hydrogen bonding, with a maximum adsorption capacity of 85.1 mg g-TS-1 for a commercial titanium silicate zeolite (TS-1), 112 mg g-TS-1 for thiol modified TS (TS-SH) and 164 mg g-TS-1 and sulfonic acid functionalized TS (TS-SO3H) (Table 2) [26,28]. Synthetic polymers feature high surface area, unique pore structure, and hydrophobicity and also demonstrated high affinity for 1,4-dioxane, especially at low concentration ranges (µg L⁻¹ levels). The adsorption capacity of Ambersorb^™ 560, a carbonaceous adsorbent, was approximately an order of magnitude higher than GAC when the equilibrium concentration of 1,4-dioxane fell below 100 µg L⁻¹ [29]. Moreover, synthesized resorcinarene cavitand polymers had reported 86% removal of 1,4-dioxane with an initial concentration of 100 µg L⁻¹, making it a promising adsorbent for 1,4-dioxane remediation [30].

Table 2. Adsorption and desorption characteristics of proposed materials for 1,4-dioxane removal.

Adsorbent	qmax (mg-1,4-Dioxane g-Adsorbent−1)	Desorption	Reference
Norit 1240 GAC	59.56	50% desorption when rising with ammonia mineral salt medium	[24]
Sawdust GAC	0.410	Not tested	[25]
ZSM-5 zeolite	22.44 to 107.36	Not tested	[27]
Titanium silicate (TS-1)	85.1	60% desorption when rising with mineral salt medium	[26]
Thiol-functionalized titanium silicate (TS-SH)	112	Quick desorption with 1 M HNO3	[28]
Sulfonic acid functionalized titanium silicate (TS-SO3H)	164	Quick desorption with 1 M HNO3	[28]
AmbersorbTM 560 polymer	~200	Not tested	[29]
Resorcinarene cavitand polymers	N/A	Not tested	[30]

Table 2. Adsorption and desorption characteristics of proposed materials for 1,4-dioxane removal.

Adsorbent	qmax (mg-1,4-Dioxane g-Adsorbent−1)	Desorption	Reference
Norit 1240 GAC	59.56	50% desorption when rising with ammonia mineral salt medium	[24]
Sawdust GAC	0.410	Not tested	[25]
ZSM-5 zeolite	22.44 to 107.36	Not tested	[27]
Titanium silicate (TS-1)	85.1	60% desorption when rising with mineral salt medium	[26]
Thiol-functionalized titanium silicate (TS-SH)	112	Quick desorption with 1 M HNO3	[28]
Sulfonic acid functionalized titanium silicate (TS-SO3H)	164	Quick desorption with 1 M HNO3	[28]
AmbersorbTM 560 polymer	~200	Not tested	[29]
Resorcinarene cavitand polymers	N/A	Not tested	[30]

The main 1,4-dioxane removal mechanism via adsorption was reported to be non-chemical interactions such as hydrophobic interactions, hydrogen bonding, and van de Waals interactions of 1,4-dioxane and the adsorbent [25,26,27,28]. Since 1,4-dioxane molecules are amphiphilic with both alkyl and oxygen groups, both hydrophobic and hydrophilic interactions may influence their adsorption behavior with different adsorbents [26]. For example,1,4-dioxane removal by TS-1 was mainly driven by hydrophobic interactions, and the adsorption capacity of 1,4-dioxane by TS-1 changed little when pH increased from neutral to 11 [26]. On the other hand, the pH-driven intermolecular hydrogen bonds between 1,4-dioxane and hydroxyl and sulfonic acid moieties resulted in significant changes in 1,4-dioxane adsorption capacity of TS-SH and TS-SO₃H over the pH range of 2–10, with an optimum pH condition of 8.6 and 7.1, respectively [28]. Nevertheless, the adsorption process does not break down or detoxify 1,4-dioxane but only transfers the contaminant from the aqueous to the solid phase, and the inherent reversibility of adsorption presents the possibility of recontamination when the influent water concentration fluctuates and falls below the adsorption equilibrium concentration [24,27].

3.2. 1,4-Dioxane Removal by Advanced Oxidation Processes

While conventional oxidation process (i.e., ozone or chlorine) has been proven to be ineffective in 1,4-dioxane degradation [31,32,33], advanced oxidation processes (AOPs) have been demonstrated to be a feasible and effective option to degrade 1,4-dioxane completely. AOPs were first proposed in the 1980s for potable water treatment and had been broadly applied in organic pollutants removal from various water sources [34,35]. AOPs are composed of two main steps: (i) in-situ generation of highly oxidative radicals (e.g., hydroxyl radicals (OH·) and sulfate radicals (S₂O₈⁻·)) and (ii) the reaction of the oxidative radicals with targeted contaminants [35]. AOPs came in different configurations based on different methods of activation and generation of free radicals, and OH· radical is one of the commonly used radical species (Figure 1). Current available OH·-utilizing AOPs for the treatment of 1,4-dioxane include the use of UV-based AOPs such as ultraviolet radiation (UV coupled with peroxide (H₂O₂) (UV/H₂O₂) [36] and UV coupled with ozone (O₃) (UV/O₃) [37]; ozone-based AOPs such as O₃/H₂O₂ [32,33,37]; catalytic AOPs such as Fenton process (Fe²⁺/H₂O₂) [38] and photo-Fenton process [39] (Figure 1). Among different AOP configurations, UV/H₂O₂ system has been widely used for the treatment of 1,4-dioxane-contaminated water from bench to full scale [2,36,40,41]. With the presence of 10 mg L⁻¹ H₂O₂, 90% removal of 100 µg L⁻¹ 1,4-dioxane was achieved in 3.5 min in bench-scale tests [41]. In the pilot test, ~80% removal of 15 µg L⁻¹ 1,4-dioxane was reported with a short UV radiation time of 26 s, and a relatively low H₂O₂ dose of 6.2 mg L⁻¹ [40]. A full-scale UV/H₂O₂ system installed at Pall-Gelman Sciences site in Ann Arbor, Michigan reported a 99.95% 1,4-dioxane removal treating 5000 gallons of contaminated water per day [2]. The decay of 1,4-dioxane by UV/H₂O₂ followed first-order kinetics, and the rate constant was reported to be 0.3–8.7 × 10⁻³ s⁻¹ with different reactor configurations, experimental conditions (H₂O₂ dose and UV wavelength), and water qualities [36,41,42]. During UV/H₂O₂ treatment, 1,4-dioxane was oxidized to intermediates such as aldehydes (formaldehyde, acetaldehyde, and glyoxal), organic acids (formic, acetic, and oxalic acid), and eventually carbon dioxide (CO₂) [36,41,42,43]. AOPs utilizing other oxidation reagents, such as UV/S₂O₈²⁻ [43,44], UV/HOCl [44], UV/NH₂Cl [45], and titanium-based photocatalysis [46,47] had also been extensively studied for the breakdown of 1,4-dioxane. Meanwhile, other emerging AOPs, such as electrochemical oxidation [48,49,50], plasma-based AOPs [51,52,53], and electron beam [54,55] have been reported to be effective (>99%) in 1,4-dioxane removal. These techniques generate hydroxyl radicals through energy inputs such as electricity, accelerated electrons, and electro-hydraulic discharge initiated by light irradiation (Figure 1).

AOPs are often favored for the treatment of 1,4-dioxane-contaminated water due to the short exposure timeline (typically several seconds to a few hours) and high removal efficiency (>90%). However, the free radicals produced during AOPs are non-selective and react with other chemicals, such as natural organic matter (NOM) and nitrate, which may negatively affect the 1,4-dioxane degradation efficiency [41]. In addition, the oxidation intermediates (e.g., aldehydes, organic acids, and esters) generated during AOPs of 1,4-dioxane and the background NOM may contribute to the formation of disinfection byproducts (DBPs) during the disinfection step [40,56,57]. Some of the AOP byproducts, such as formaldehyde and glycolaldehyde, may be more geno- and cyto-toxic than 1,4-dioxane [58,59,60]. Furthermore, a high concentration of oxidants (e.g., 8–10 mg L⁻¹ of H₂O₂) is often required to generate sufficient free radicals to achieve complete degradation of 1,4-dioxane. Therefore, a subsequent step such as GAC is required to consume the residual oxidant as well as the AOP byproducts [61,62,63].

4. Biological Treatment for 1,4-Dioxane Removal

Compared with chemical approaches, biodegradation treatment is more cost-effective and environmental-friendly, and may be a promising approach for 1,4-dioxane removal [64]. Over the past decades, significant effort has been made in the field of 1,4-dioxane bioremediation. More than 30 pure 1,4-dioxane degrading strains and microbial communities had been obtained from contaminated or uncontaminated environments. The functional microbial species and the 1,4-dioxane biodegradation pathways were well-studied [65,66,67,68,69,70]. In addition, various bioreactors employing pure 1,4-dioxane degrading strains or microbial communities were developed to remove 1,4-dioxane from contaminated surface water, groundwater, industrial wastewater, and landfill leachate [15,71,72,73,74]. Compared with energy-intensive physicochemical treatment processes such as AOP/UV, biological treatment is more economically efficient and can result in the complete degradation of 1,4-dioxane to CO₂.

4.1. Microbiology of 1,4-Dioxane Degrading Pure Strains and Microbial Communities

4.1.1. Aerobic 1,4-Dioxane Biodegradation

A number of microbial species and microbial communities were reported to degrade 1,4-dioxane through either metabolism (growth-supporting) or co-metabolism (non-growth supporting) under aerobic conditions (

Table 3. Pure strains and microbial communities that can degrade 1,4-dioxane via metabolism or co-metabolism.

Strain	Induced Enzyme	Co Substrate	Biodegradation Rate a	qmax (mg-Dioxane h−1 mg-Protein −1)	Ks (mg L−1)	Enrichment Source	Reference
Co-metabolic strains
Pseudonocardia K1	THF MO	THF	0.26 ± 0.013 mg hr−1 mg-protein−1				[76]
Rhodococcus RR1	N/A	Toluene	0.38 ± 0.03 mg hr−1 mg-protein−1
Methylosinus trichosporium OB3b	sMMO	Methane	0.38 ± 0.02 mg hr−1 mg-protein−1
Mycobacterium vaccae JOB5	Propane MO	Propane	0.40 ± 0.06 mg hr−1 mg-protein−1
Pseudomonas mendocina KR1	toluene-4-MO	Toluene	0.37 ± 0.04 mg hr−1 mg-protein−1
Ralstonia pickettii PKO1	toluene-p-MO	Toluene	0.31 ± 0.007 mg hr−1 mg-protein−1
Burkholderia cepacia G4	toluene-2-MO	Toluene	0.1 ± 0.006 mg hr−1 mg-protein−1
Pseudonocardia sp. strain ENV478	N/A	THF	0.008 mg hr−1 mg-protein−1				[77]
Mycobacterium sp. strain ENV421	Propane MO	Propane	N/A				[80]
Azoarcus sp. DD4	Toluene MO	Toluene	1.82 mg L−1 day−1				[79]
Graphium sp. (ATCC 58400) (fungus)	MO	THF	19 ± 10.5 mg hr−1 mg-protein−1				[81]
Metabolic strains
Pseudonocardia dioxanivorans CB1190	1,4-dioxane MO		0.19 ± 0.007 mg hr−1 mg-protein−1	1.1 ± 0.0008	160 ± 44	1,4-dioxane-contaminated industrial sludge	[76]
P. benzenivorans B5	N/A		0.01 ± 0.003 mg hr−1 mg-protein−1	0.1 ± 0.006	330 ± 82	Contaminated soil
Mycobacterium sp. PH-06	MO		2.5 mg L−1 h−1	N/A	78 ± 10	Contaminated sediment	[66]
Acinetobacter baumannii DD1	MO		2.38 mg L−1 h−1	N/A	N/A	Activated sludge	[82]
Afipia sp. D1	N/A		0.263 mg hr−1 mg-protein−1	0.263	25.8	Drainage of a chemical factory	[67]
Mycobacterium sp. D6			0.139 mg hr−1 mg-protein−1	0.139	20.6
Mycobacterium sp. D11			0.052 mg hr−1 mg-protein−1	0.052	69.8
Pseudonocardia sp. D17			0.096 mg hr−1 mg-protein−1	0.096	59.7
Xanthobacter flavus DT8	MO		Equivalent to CB1190	N/A	17.5	Activated sludge of pharmaceutical plant	[83]
Rhodococcus aetherivorans JCM 14343			0.0073 mg hr−1 mg-protein−1	0.0073	59.2	N/A	[84]
Pseudonocardia carboxydivorans. RM-31	N/A		31.6 mg L−1 h−1	N/A	N/A	Seawater	[85]
Ancylobacter phlymorphus ZM13	Toluene MO		N/A	N/A	N/A		[86]
Microbial community
Consortium CH1	Toluene MO			2.04	N/A	Activated sludge	[86]
Mixed culture	N/A			0.019	11.08	Activated sludge	[73]
Consortium A	Propane MO			0.297 ± 0.0075 (at 500 mg L−1)	N/A	Uncontaminated soil	[70]
Consortium B	Propane MO			0.236 ± 0.0029 (at 500 mg L−1)	N/A	Uncontaminated soil	[70]
Enrichment culture-FS	N/A			0.037	93.9	Forest soil	[87]
Enrichment culture-AS	N/A			0.078	181.3	Activated sludge	[87]
Soil–enrichment	1,4-dioxane MO			0.044 ± 0.001	25 ± 1.60	Uncontaminated soil	[88]

Notes: MO: monooxygenase. ^a: For consistency of dimensions, the following conversions are used: 1 g protein = 0.5 g VSS = 0.4 g TSS = 0.43 g COD [76]. Values are given as means the standard deviations where applicable. N/A, not applicable.

). Co-metabolism is the biotransformation of non-growth-supporting substrates in the obligate presence of a growth-supporting substrate [75]. Various primary substrates had been identified for effective 1,4-dioxane removal, including tetrahydrofuran, propane, methane, ethane, isobutane, and toluene, and 1,4-dioxane is co-metabolically biodegraded by monooxygenase induced during the biodegradation of these primary substrates [76,77,78]. Such strains include Pseudonocardia K1 growing on tetrahydrofuran [76], Pseudonocardia sp. strains ENV478 growing on tetrahydrofuran [77], Azoarcus sp. DD4 growing on toluene [79] (Table 3). The biodegradation rates of co-metabolic biodegradation are at similar levels to the metabolic 1,4-dioxane biodegradation by the well-studied Pseudonocardia dioxanivorans sp. nov. CB1190 (0.1 to 0.40 mg h⁻¹ mg-protein⁻¹). The co-metabolic degradation of 1,4-dioxane as well as the co-contaminant 1,1-DCE by Azoarcus sp. DD4 was reported when propane was added as the primary substrate, and a toluene monooxygenase was confirmed as the key enzyme for the degradation of propane, 1,4-dioxane as well as 1,1-DCE [79]. The co-metabolic biodegradation of 1,4-dioxane requires the addition of the primary substrates, and for some strains such as Pseudonocardia sp. strain ENV478, the biodegradation performance may be negatively affected by the accumulation of 1,4-dioxane biodegradation intermediate (2-hydroxyethoxyacetic acid (HEAA), however, it has become a promising approach for treating 1,4-dioxane at environmentally relevant concentrations (µg L⁻¹ levels) to meet the stringent 1,4-dioxane guidelines and standards, as the growth of the microorganism and the enzyme induction were supported by the primary substrates rather than the trace level of 1,4-dioxane.

Metabolic 1,4-dioxane degraders can utilize 1,4-dioxane as the sole carbon and energy source. Pseudonocardia dioxanivorans sp. nov. CB1190 (CB1190) was isolated from industrial wastewater sludge and had been comprehensively studied. 1,4-Dioxane can be biodegraded by CB1190 at a biodegradation rate of 0.19 ± 0.007 mg h⁻¹ mg-protein⁻¹ [76]. Further studies revealed the biodegradation pathway as well as the initiation of 1,4-dioxane biodegradation by a multicomponent dioxane monooxygenase (dxmADBC) [89]. Since then, numerous studies reported the isolation of 1,4-dioxane metabolizers with alternative pathways and distinct biodegradation and growth kinetics (Table 3). Most of the current reported 1,4-dioxane degrading bacteria were enriched from 1,4-dioxane contaminated sources, such as contaminated industrial sludge (CB1190), contaminated soil (P. benzenivorans B5), contaminated sediment (Mycobacterium sp. PH-06), activated sludge (Acinetobacter baumannii DD1and Xanthobacter flavus DT8) and seawater (Pseudonocardia carboxydivorans. RM-31). The biodegradation rate of metabolic 1,4-dioxane biodegradation varies from 0.01 mg h⁻¹ mg-protein⁻¹ of P. benzenivorans B5 to 0.263 mg h⁻¹ mg-protein⁻¹ of Afipia sp. D1 (Table 3). Sludge-enriched Xanthobacter flavus DT8 exhibited a low half-saturation constant of 17.5 mg L⁻¹ and a high specific growth rate of 0.15 h⁻¹, showing the feasibility of large-scale cell cultivation [83]. Distinct Monod maximum 1,4-dioxane biodegradation rate (q_max) (0.01–1.1 mg-dioxane h⁻¹ mg-protein⁻¹) and half-saturation constant (K_s) (17.5–330 mg L⁻¹) were observed for the reported metabolic 1,4-dioxane degraders. The high specific dioxane degradation rate (0.263 mg-dioxane h⁻¹ mg-protein⁻¹) and low half-saturation constant (25.8 mg L⁻¹) of Afipia sp. D1 suggested its high activity and affinity toward 1,4-dioxane degradation [67]. The metabolic 1,4-dioxane degrading strains are found to be soluble di-iron monooxygenase (SDMO) expressing strains, and are commonly found in genera of Pseudonocardia, Mycobacterium, Rhodococcus, Xanthobacter, and Pseudomonas [76,83,84,90].

In addition, the majority of 1,4-dioxane-degrading pure strains are bacteria. Only two fungi species have been reported to be able to degrade 1,4-dioxane. Cordyceps sinensis was enriched from garden soil and was able to metabolically degrade 0.09 M (7.92 g L⁻¹) 1,4-dioxane. Rapid 1,4-dioxane degradation efficiency (90% removal in 3 days) was observed when cultivated at its optimal growth 1,4-dioxane concentration of 0.034 M (3 g L⁻¹), however, 1,4-dioxane degradation at lower concentrations of 0.0125 × 10⁻³ M (1.1 mg L⁻¹) was incomplete [91]. When growing on tetrahydrofuran or propane as the primary substrate, Graphium sp. (ATCC 58400) was identified to co-metabolically degrade 1,4-dioxane. 50% removal was observed in 12 h with propane as the primary growth substrate and an initial 1,4-dioxane concentration of 17.6 mg L⁻¹). The specific biodegradation rates of 1,4-dioxane were reported to be 4 ± 1 and 9 ± 5 nmol substrate min⁻¹ mg dry weight⁻¹ (0.021 ± 0.005 and 0.048 ± 0.026 mg-1,4-dioxane h⁻¹ mg dry weight⁻¹) [81].

The enriched microbial communities capable of metabolic 1,4-dioxane biodegradation were obtained from various sources, such as activated sludge, contaminated/uncontaminated soil, and landfill leachate [70,86,88,92,93,94,95,96,97]. Microbial communities that were reported to be capable of 1,4-dioxane biodegradation are listed in Table 3. 1,4-Dioxane biodegradation by microbial communities was found to be initiated by monooxygenases such as toluene MO of Consortium CH1 [86], propane MO by soil-enriched Consortium A and B [70], and dioxane MO by a soil-enriched microbial community [88] (Table 3). The q_max and K_s values of current reported microbial communities were close to the q_max and K_s values of pure strains. The toluene MO expressing consortium CH1 reported a maximum 1,4-dioxane biodegradation rate of 2.04 mg h⁻¹ mg-protein⁻¹, which was almost twice the well-studied pure strain CB1190 [86]. Smaller K_s of 11.08 and 25 ± 1.6 mg L⁻¹ were reported in two 1,4-dioxane degrading microbial communities, respectively, showing a great affinity toward 1,4-dioxane biodegradation [73,88]. While 1,4-dioxane biodegradation was initiated by 1,4-dioxane degraders in the enrichment, the coexistence of non-1,4-dioxane degraders may assist the mineralization of intermediates and achieve a faster 1,4-dioxane degradation rate [86]. Also, the application of microbial communities for bioremediation may be more appealing since most of the bioreactors for water treatment are based on mixed microbial communities [52].

4.1.2. Functional Enzymes for 1,4-Dioxane Biodegradation

1,4-dioxane biodegradation is known to be initialized by soluble di-iron monooxygenase (SDMO). Currently reported monooxygenases responsible for metabolic 1,4-dioxane biodegradation include the dioxane monooxygenase (dxmADBC) of CB1190, propane monooxygenase (prmABCD) of PH-06 and toluene monooxygenase (tmoABCDEF) of Ancylobacter polymorphus ZM13, [76,86,98]. Bacterial strains growing on other substrates, such as hydrocarbons, cyclic ethers, and aromatic hydrocarbons, may also induce SDMO that may initiate 1,4-dioxane degradation. Such SDMOs include soluble methane monooxygenase (sMMO) of Methylosinus trichosporium OB3b [76], tetrahydrofuran (THF) monooxygenase of Pseudonocardia sp. ENV478 [77], propane monooxygenase of Mycobacterium vaccae JOB5 [76] and toluene monooxygenase of Azoacus sp. strain DD4 [69]. The biodegradation pathways of several species had been well studied (Figure 2), with 2-hydroxy-1,4-dioxane, 2-hydroxyethoxyacetic acid (HEAA), ethylene glycol, glycolate, and oxalate as major biodegradation intermediates [86,89,98]. For some co-metabolic strains such as Pseudonocardia sp. strain ENV478, the biodegradation process ceased at the production and accumulation of HEAA as the end product [77].

Time-course evaluation of gene expression of CB1190 confirmed the upregulation of the dxmB gene of dioxane monooxygenase (DXMO) in response to dioxane exposure [99]. Currently, the gene clusters dxmA and dxmB of DXMO gene and the gene (thmA) encodes for its homolog tetrahydrofuran monooxygenase (THM) of Pseudonocardia tetrahydrofuranoxydans K1, are currently used as biomarkers to evaluate the 1,4-dioxane degradation ability of pure strains and microbial communities as well as the 1,4-dioxane degradation potentials in contaminated sites [88,100,101]. The prmA gene encodes for propane monooxygenase clusters (PRM) in Mycobacterium dioxanotrophicus PH-06 was upregulated in the presence of dioxane, and its abundance was significantly correlated with dioxane degradation activity, and can be a potential biomarker to evaluate the extent of 1,4-dioxane biodegradation [90].

4.1.3. Effect of Co-Contaminants on 1,4-Dioxane Biodegradation

1,4-Dioxane is frequently used as a co-stabilizer with chlorinated solvents in industrial practice. Many studies have reported the common occurrence of trichloroethylene (TCE) and 1,4-dioxane at contaminated sites [5,102,103]. The inhibition of chlorinated solvents on 1,4-dioxane biodegradation had been previously identified following the order of VC > 1,1-DCE > cis-DCE > TCE > TCA [101,104,105,106]. The inhibition mechanism of TCA and 1,1-DCE on CB1190 had been determined as non-competitive inhibition, with an inhibition constant (K_I) of 1.2 ± 1.0 μM for TCA and 3.3 ± 2.9 μM for 1,1-DCE [104]. The inhibition of chlorinated solvents on 1,4-dioxane biodegradation was attributed to delayed ATP production and down-regulation of both DXMO and ALDH genes. 1,1-DCE concentration as low as 0.5 mg L⁻¹ may cause 40% decrease of the 1,4-dioxane removal efficiency by CB1190 [101] and 1 mg L⁻¹ 1,1-DCE may completely inhibit 1,4-dioxane biodegradation by soil-enriched microbial community [88]. Less inhibition of chlorinated solvent was observed when CB1190 was grown in attached forms. For example, in a flow-through column bioaugmented with CB1190, 1,4-dioxane removal efficiency decreased by 9% when influent 1,1-DCE concentration of 5 mg L⁻¹ [105] while almost complete inhibition of 1,4-dioxane degradation was observed with the same 1,1-DCE concentration in batch studies [101].

Though no TCE or 1,1-DCE removal had been observed during the 1,4-dioxane biodegradation by CB1190, the aerobic biodegradation of cis-DCE by CB1190 was reported, with the biodegradation rate of approximately 200 µg L⁻¹ day⁻¹ [107]. While the 1,4-dioxane biodegradation rate (22.3 ± 5 µg L⁻¹ h⁻¹) with the presence of VC (4000 µg L⁻¹) was approximately 50% compared to the level with the presence of similar concentrations of 1,1-DCE (5000 µg L⁻¹). Meanwhile, CB1190 was reported to metabolically degrade VC through alkaline monooxygenase [106]. CB1190 can aerobically biodegrade 40 µg L⁻¹ of VC to below the detection limit (25 µg L⁻¹) in 1 day at a biodegradation rate of 1.6 ± 0.05 µg L⁻¹ h⁻¹.

4.1.4. Anaerobic 1,4-Dioxane Biodegradation

1,4-Dioxane is generally considered resistant to biodegradation under anaerobic conditions. To date, limited research has reported the feasibility of anaerobic 1,4-dioxane biodegradation. An iron-reducing bacterium enriched from wastewater treatment plant sludge was reported to metabolically degrade 1,4-dioxane under anaerobic conditions. When the bacterium was supplied with Fe(III)-EDTA as the electron donor and humic acid as an electron shuttling agent, 90% removal of 13 mg L⁻¹ of 1,4-dioxane was achieved in 40 days [108]. In another study, biodegradation of up to 70 mg L⁻¹ of TCA and 1,4-dioxane by an iron-reducing microbial community was reported to follow first order with the apparent first-order kinetic coefficient of 0.057 ± 0.003 h⁻¹ for TCA and 0.012 ± 0.003 h⁻¹ for 1,4-dioxane, respectively [109]. The biodegradation ability was further promoted by the addition of zero-valent iron with the apparent first-order kinetic coefficient of 0.175 ± 0.009 h⁻¹ for TCA and 0.272 ± 0.002 h⁻¹ for 1,4-dioxane, respectively. Another recent study examined the anaerobic 1,4-dioxane biodegradation potential of sediment samples collected from various environments (uncontaminated agricultural soil, uncontaminated river sediments, and contaminated soil) with different electron acceptor (i.e., iron, sulfate, nitrate) amendments. During the 1-year incubation period, 1,4-dioxane biodegradation was observed in the nitrate-amended treatment (16 ± 16% removal) and the treatment with no electron acceptor amendment (22 ± 19% removal), suggesting the potential of 1,4-dioxane biodegradation under anaerobic and high reducing conditions [110]. However, no anaerobic biodegradation pathway of 1,4-dioxane has been discovered so far.

4.2. Biotechnologies for 1,4-Dioxane Removal

4.2.1. Suspended Growth Bioreactors for 1,4-Dioxane Removal

The application of biological treatment technologies for 1,4-dioxane removal had been explored using different types of bioreactors (

Table 4. Summary of reported bioreactor for 1,4-dioxane removal.

Reactor Type	Dimension	Water Source	Microbial	HRT	1,4-Dioxane Loading Rate	Influent 1,4-Dioxane Concentration	Removal Efficiency	Reference
Packed soil flow-through column	2.5 cm ID × 10.5 cm H (10 cm packing height)	Contaminated groundwater	Pseudonocardia dioxanivorans CB1190	41.9–80.8 h	0.043–0.144 mg-1,4-dioxane d−1	3–10 mg L−1	Up to 99% influent concentration of 10 mg L−1	[105]
Packed sand filtration column	5 cm ID × 120 cm H (100 cm packing height)	Contaminated groundwater	Pseudonocardia dioxanivorans CB1190	66–277 h of EBCT		0.2–5 mg L−1	34–92%	[113]
Tire chips packed up-flow biological aerated filter	20 cm ID × 79.5 cm H (27 cm packing height)	1,4-dioxane containing wastewater	Activated sludge	5–7 h		17.8–65.6 mg L−1	54.7–83.4%	[114]
Biological activated filter	Sequential column of 2.54 cm ID × 32 cm packing height	Filtered water from water treatment plant	Co-metabolic microbial community enriched from river basin sample	7.5–30 min of EBCT		8.9–11.7 µg L−1	65–94%	[74]
Continuous Stirred tank reactor (CSTR)	15 L	Industrial wastewater	Microbial community enriched from activated sludge	10–40 h		200 mg L−1	81.6–98.6%	[73]
Plug Flow reactor (PFR)	11 L	Industrial wastewater	Microbial community enriched from activated sludge	10–40 h		200 mg L−1	96.2–99.8%	[73]
Moving bed bioreactor with glycol gel carriers	1000 mL Carrier packing ratio 15%	Synthetic industrial wastewater	Afipia sp. D1	16–24 h	0.4–0.6 kg 1,4-dioxane m−3 d−1	~400 mg L−1	99%	[115]
Tubular carrier/polyurethane carrier	1000 mL	Synthetic wastewater	Pseudonocardia sp. D17	2.4–20 h		5–50 mg L−1	90–99	[116]
PEG gel beads (15% packing ratio)	1440 mL	Synthetic wastewater	Pseudonocardia sp. D17	12–60 h		20–100 mg L−1	70–99	[117]
PEG gel carriers	700 mL	Synthetic wastewater	Sludge-enriched 1,4-dioxane degrading consortium	3–6 h		5–40 mg L−1	79–99.5	[118]
Immobilized gel-carrier bioreactor	120 L	Industrial wastewater	Afipia sp. D1	40 h	0.09–0.47 kg-dioxane m−3 d−1	2–670 mg L−1		[71]
Moving bed biofilm reactor	1050 mL	Basal salt medium (BSM)	Microbial samples from landfill leachate treatment facility	0.5–2 d	5.14–20.8 mg L−1 d−1	10 mg L−1	69.4–97.9% at different conditions	[72]
Semicontinuous stirred tank reactor	5 L	Industrial wastewater	Rhodanobacter AYS5	6 h		263 mg L−1	100% in 6 days	[111]
Sequential batch membrane bioreactor	8.4 L	Synthetic wastewater	activated sludge	7.3 h		100–500 mg L−1	27.36–94.3% with different amounts of acetate addition	[112]
Trickling filter packed with ceramic saddles	1.3 cm × 1.1 m	25% mineral medium L amended with THF and 1,4-dioxane	Co-metabolic microbial community enriched from a contaminated aquifer	14.4 min		0.99–1.51 mg L−1	>93%	[119]

). Continuous stirred tank reactor (CSTR) (Figure 3a) and plug flow reactor (PFR) (Figure 3b) had been applied for the treatment of industrial wastewater with high 1,4-dioxane concentrations (200 mg L⁻¹) [73]. The reactors were inoculated with 1,4-dioxane degrading microbial communities enriched from activated sludge. High 1,4-dioxane removal efficiency (81.6–98.6% in CSTR and 96.2–99.8% in PFR) was achieved under various hydraulic retention times (10–40 h). Lab-scale semi-continuous stirred tank reactor inoculated with a metabolic strain Rhodanobacter AYS5 achieved 100% removal of 1,4-dioxane in 6 days when a high influent 1,4-dioxane concentration of 263 mg L⁻¹ [111]. In a recent study of a sequential batch membrane bioreactor inoculated with sludge-enriched 1,4-dioxane degrading microbial community, the 1,4-dioxane removal efficiency increased from 31.1 ± 3.7% to 87.5 ± 6.8% when the amended acetate to 1,4-dioxane ratio increased from 0 to 4 [112].

4.2.2. Immobilized Cell Bioreactors for 1,4-Dioxane Removal

Immobilized cell bioreactors (Figure 3c) employing pure 1,4-dioxane degrading strains/microbial community were also developed to remove 1,4-dioxane from industrial wastewater with high 1,4-dioxane concentration (5–670 mg/L) [71,72,115,116,117,118]. These systems were designed to maintain a high bacterial cell density for adequate degradation activities of the 1,4-dioxane degraders, and subsequently a high 1,4-dioxane removal performance. A series of studies reported the 1,4-dioxane removal performance by immobilized cell bioreactors using different carrier materials (i.e., silicon rubber carrier, polyethylene carrier, polyurethane carrier, and polyethylene glycol (PEG) gel carriers) and different 1,4-dioxane degrading species or microbial communities (i.e., Pseudonocardia sp. D17, Afipia sp. D1, sludge, and landfill leachate enriched microbial communities) [72,115,116,117,118].

For example, polyethylene glycol (PEG) gel was used as the carrier for 1,4-dioxane degrading species, and effective 1,4-dioxane removal was achieved by various scales of the bioreactor (700 mL to 120 L) packed with this biocarrier [71,115,115,117,118]. PEG cell reactors immobilized with sludge-enriched 1,4-dioxane-degrading consortium showed effective 1,4-dioxane removal (99.5%) with a short HRT (6 h) [118]. The performance of PEG gel carriers immobilized with Afipia sp. D1 showed more than 99% removal of 1,4-dioxane under both bench and pilot scale conditions [71,116] The impact of nutrients level on the 1,4-dioxane removal in an immobilized cell bioreactor utilizing PEG gel carrier immobilized with Pseudonocardia sp. D17 showed an effective removal of 1,4-dioxane from 5–50 mg L⁻¹ to 0.49 mg L⁻¹ even at low temperatures (15 °C) [116,117]. The feed of the essential nutrients requires the ratio of 1,4-dioxane: N:P:S = 20:2.0:0.4:0.1 for efficient treatment.

The performance of the lab-scale bioreactor utilizing different carrier materials (silicon rubber carrier, polyethylene carrier, and polyurethane carrier) immobilized with 1,4-dioxane degrading microbial community enriched from leachate treatment facility was evaluated for the treatment of synthetic wastewater containing 10 mg/L 1,4-dioxane under hydraulic retention times (0.5–2 days) and aeration strength (0.1–1 volume of air per volume of water per minute (vvm)) [72]. Low effluent1,4-dioxane levels (0.5 mg L⁻¹) were observed in bioreactors packed with different carrier types. Though sufficient aerobic condition (DO > 6 mg L⁻¹) was maintained under different aeration strengths, 1,4-dioxane removal performance of the bioreactor employed with silicone rubber carrier decreased drastically from 95% to 35% when aeration strength reduced to 0.1 vvm due to the accumulation of chloroform as an intermediate. Under a short HRT of 0.5 days (1,4-dioxane loading rate of 20.8 mg L⁻¹ d⁻¹), the 1,4-dioxane removal efficiency of silicon rubber carrier and polyurethane carrier bioreactor decreased from ~95% to 82.7 ± 1.3% and 67.1 ± 3.5%, respectively, and the carrier characteristics and control of operating conditions may be crucial to ensure a stable and effective 1,4-dioxane removal by the immobilized cell bioreactors.

4.2.3. Biofiltration System for 1,4-Dioxane Removal

Biofiltration systems (Figure 3d) employed with metabolic or co-metabolic microbes had been developed with distinct configurations and operation parameters [74,105,113,114,119]. Gravity-flow mini soil columns (height 10 cm, diameter 2.5 cm) bioaugmented with strain Pseudonocardia dioxanivorans CB1190 showed nearly complete degradation of 3–10 mg L⁻¹ 1,4-dioxane under low hydraulic loading (0.01–0.02 mL min⁻¹) [105]. Though the 1,4-dioxane removal was partially inhibited by the presence of 1,1-DCE and Cu²⁺ ions, the inhibition was less severe compared with batch experiments with the same level of co-contaminants. 1,4-Dioxane biodegradation by the inoculation of CB1190 was also investigated in larger up-flow columns [113]. Significant 1,4-dioxane removal was observed only at the inoculation point with a high 1,4-dioxane concentration (5 mg L⁻¹). The degradation was stopped after the hydraulic loading rate decreased to 0.36 cm h⁻¹ since the dissolved oxygen level dropped to below 0.5 mg L⁻¹ and eventually lost biodegradation activity. Under environmentally relevant 1,4-dioxane concentration (200 µg L⁻¹), the removal was less effective (34% removal) since a higher 1,4-dioxane concentration was needed to induce and sustain significant bacteria metabolism [113]. A sludge-seeded up-flow biological activated filter (UBAF) was packed with tire chips to treat 1,4-dioxane-contaminated wastewater [114]. The metabolic BAF showed an average removal of 71% with an average influent 1,4-dioxane concentration of 31 mg L⁻¹. The optimal EBCT and air: water ratio was found to be 8.5 h and 30:1, and 1,4-dioxane removal efficiency increased from 59.5% to 72.8%1,4-dioxane when increasing 1,4-dioxane loading rate from 427to 986 mg-dioxane d⁻¹. Overall, the above metabolic columns achieved effective removal of 1,4-dioxane (>70%) when the influent concentration was at mg L⁻¹ level.

On the other hand, the application of co-metabolic 1,4-dioxane degrading strains demonstrated a better removal efficiency at environmentally relevant concentrations (µg L⁻¹ level) [74,119]. The ability of a lab-scale trickling filter (length 1.1 m, diameter 13 mm) to biodegrade 1,4-dioxane in the presence of tetrahydrofuran (THF) was investigated [119]. The trickling filter was capable to biodegrade 93–97% of 1,4-dioxane with the influent containing 0.2–1.25 mg 1,4-dioxane L⁻¹ and THF (6.0–21.0 mg L⁻¹) as the growing substrate. Pilot-scale biofiltration columns were developed to remove 1,4-dioxane at drinking water-relevant concentrations (8.9–11.7 µg L⁻¹) [74]. The gravity-driven columns were filled with granular activated carbon and butane (1% in the airflow) was added as the growing substrate. With an influent 1,4-dioxane concentration of 10 µg L⁻¹, 54 to 87% removal of 1,4-dioxane was achieved with different GAC particle sizes and the empty bed contact time (EBCT) from 7.5–30 min. The highest removal efficiency of 87 ± 5% was obtained with the largest GAC particle sizes (effective size of 1.9 mm) under the highest EBCT of 30 min since the penetration of butane can be achieved with larger GAC grains and longer contact times.

4.2.4. Biostimulation and Natural Attenuation

Bioremediation can occur naturally through natural attenuation, which relies on the intrinsic biodegradation potential of indigenous microbial populations to biodegrade the pollutant in the impacted soil [120]. Though 1,4-dioxane is considered recalcitrant to biodegradation, significant natural attenuation was confirmed in 22 monitoring sites, with a median first-order site attenuation rate coefficient of 0.26 yr⁻¹ [102]. Microcosm studies also showed a significant correlation (p < 0.05) between the abundance of thmA/dxmA genes and 1,4-dioxane degradation activity [100]. However, the analysis of currently available information on 1,4-dioxane biodegrading microorganisms, their associated molecular mechanisms of degradation, and potential biomarkers with confidence levels suggested the distribution of naturally occurring potential for dioxane bioremediation may be limited [121].

1,4-Dioxane contaminate plumes expand rapidly upon the chemical release into the aquifers [122,123]. Therefore in-situ bioremediation may be more efficient and cost-effective for large and diluted 1,4-dioxane plumes. Propane biosparging with the addition of co-metabolic 1,4-dioxane degrader Rhodococcus ruber strain ENV425 had been studied at both pilot and field scales (Figure 3e) [124,125,126]. Pilot-scale in-situ tests demonstrated an average of 74% 1,4-dioxane reduction at a contaminated site with well-distributed propane and air [125]. Pilot-scale propane biosparging system installed at Vandenberg Air Force Base had demonstrated successful 1,4-dioxane biodegradation in contaminated aquifers, with the 1,4-dioxane concentration decrease from 1090 µg L⁻¹ to <2 µg L⁻¹ near the sparging well and from 135 µg L⁻¹ to 7.3 µg L⁻¹ at a more distal well, respectively [124]. Additionally, the first full-scale propane biosparging system was installed at the same site. With the propane addition of 52 g per day per well, the system was reported to achieve an average 1,4-dioxane removal of 64.1% over the 6-month operation [126].

5. Challenges and Opportunities of 1,4-Dioxane Removal Technologies

5.1. Metabolic 1,4-Dioxane Degradation at Environmental Relevant Conditions

Current reported metabolic 1,4-dioxane degraders had demonstrated rapid and effective 1,4-dioxane biodegradation at high 1,4-dioxane concentrations (i.e., mg L⁻¹ levels,). However, few studies had demonstrated the feasibility of metabolic 1,4-dioxane degradation under environmentally relevant levels (µg L⁻¹) [70,88,127]. Metabolic 1,4-dioxane biodegradation at µg L⁻¹ levels may be challenging as such low concentrations were not sufficient to support microbial growth as well as induce functional enzymes [128]. The Monod kinetics of most currently reported metabolic 1,4-dioxane degraders feature large 1,4-dioxane half-velocity constant (K_s) ranging from 11.8–330 mg L⁻¹, which is several orders of magnitude higher than the environmentally relevant concentrations (µg L⁻¹ level). Thus, the enrichment of microorganisms with low K_s may be crucial for the successful application of metabolic 1,4-dioxane bioremediation. The observation of natural attenuation of 1,4-dioxane in contaminated plumes, the confirmation of 1,4-dioxane biodegradation in nitrite removing biofilters at μg L⁻¹ as well as the detection of biomarker genes associated with 1,4-dioxane biodegradation may have suggested the existence of such microorganism [129]. Furthermore, the kinetic characteristics of a 1,4-dioxane degrading microbial community confirmed the co-existence of 1,4-dioxane degrading microbial species with lower K_s of 0.44 mg L⁻¹ as well as 1,4-dioxane degrading microbial species with higher K_s [97]. Meanwhile, the identification of growth-supporting addictive, such as amino acids or vitamins, may be crucial to maintain rapid 1,4-dioxane biodegradation. Thiamine (vitamin B1) was identified as a primary limiting cofactor for 1,4-dioxane metabolism by Rhodococcus ruber strain 219. With the supplement of thiamine, repeated 1,4-dioxane biodegradation from 100 μg L⁻¹ to below the health advisory level of 0.35 μg L⁻¹ can be observed with a low K_s (0.015 ± 0.03 μg L⁻¹) [127]. Though the biodegradation of 1,4-dioxane at low concentrations (µg L⁻¹) had been proven feasible through both metabolic and co-metabolic pathways, AOPs remained the most effective and commonly used approach for water treatment at these environmentally relevant concentrations.

5.2. Co-Existence of Chlorinated Solvents

As 1,4-dioxane was mainly used as a stabilizer for chlorinated solvents, 1,4-dioxane contamination commonly co-occurs with the existence of chlorinated solvents, including TCE, TCA, 1,1-DCE, and cis-DCE. Data from the United States Air Force (USAF) Environmental Restoration Program Information Management System (ERPIMS) reported 93.5% (732 out of 781) of the groundwater monitoring wells that had detected 1,4-dioxane had also detected TCE and/or TCA [5]. Similarly, the investigation of >2000 sites in California revealed that, among the 194 sites detected with 1,4-dioxane, 95% of the sites were co-contaminated with one or more chlorinated solvents [122]. As the presence of chlorinated solvents can negatively affect the 1,4-dioxane biodegradation performance [101,104,105], and the elimination of co-occurring chlorinated solvents may be essential to achieving efficient biodegradation of 1,4-dioxane. Bench-scale studies have demonstrated the feasibility of sequential anaerobic/aerobic bioaugmentation to degrade TCE anaerobically and 1,4-dioxane aerobically [103,107]. However, both experiments were conducted in batch tests in a rather short term (<7 days), and the effectiveness of long-term sequential anaerobic/aerobic biodegradation had not been investigated. Meanwhile, the survival of 1,4-dioxane degraders under anaerobic conditions as well as their quick recovery after being resupplied with oxygen may be crucial and challenging to maintain an effective 1,4-dioxane removal [88]. Therefore, the development of an effective and feasible treatment strategy for the removal of both chlorinated solvents and 1,4-dioxane is still of urgent need.

5.3. Combined Process for 1,4-Dioxane Removal

Treatment that combines the biodegradation process with the physicochemical process had been developed aiming at enhanced 1,4-dioxane removal as well as reducing overall operating costs. 1,4-Dioxane removal efficiency by conventional ozonation followed with BAF process was reported to increase to ~50%, while the conventional ozonation and BAF alone only removed 6–11% and 25% of 1,4-dioxane, respectively [130]. Furthermore, 1,4-dioxane removal by filtration processes utilizing materials with adsorptive characteristics could be enhanced with the bioaugmentation of 1,4-dioxane degrading microorganisms [24,26,27]. Batch adsorption experiments of zeolite bioaugmented with CB1190 showed an enhanced 1,4-dioxane removal to below the equilibrium concentration of 0.18 mg L⁻¹ [27]. Also, TCE and cis-DCE were preferentially adsorbed on zeolite over 1,4-dioxane, and thus reduced the inhibition effect of chlorinated solvents on 1,4-dioxane biodegradation. Bioaugmented Norit 1240 GAC achieved a greater 1,4-dioxane removal (95–98% reduction) compared with the abiotic GAC (85–89% reduction) [24]. The electrochemical oxidation process utilizing Ti/IrO₂−Ta₂O₅ mesh electrodes in flow-through reactors coupled with bioaugmentation of CB1190 showed rapid 1,4-dioxane removal from up to 100 mg L⁻¹ to below 3 µg L⁻¹ while reducing the overall treatment cost by 1 order of magnitude, comparing with direct electrochemical 1,4-dioxane oxidation processes [49,131]. Though 1,4-dioxane was partially removed (~50%) by an electrochemical process operated at low anode potentials, the electrochemical removal of chlorinated solvents (TCE and 1,1-DCE), as well as the anodic O₂ generation, provided an advantageous environmental for aerobic biodegradation process [49,131]. Non-thermal plasma reactor combined with an aerobic bioreactor process had shown a rapid and effective 1,4-dioxane degradation by plasma and byproducts removal by biodegradation, with the energy efficiency increased to 5 times than that of a water film plasma reactor [52].

6. Conclusions

The toxicity and recalcitrance of 1,4-dioxane resulted in wide detection of this chemical in aqueous environments and increasing attention on its removal from both drinking water and wastewater. While AOPs are the most effective and readily available approaches for the full-scale treatment of 1,4-dioxane-contaminated water, special attention should be paid to the byproduct generation and the increasing DBP formation potentials. Studies utilizing both traditional and modified adsorbent materials showed a temporal removal of 1,4-dioxane. However, additional disposal and regeneration procedures are required. Effective 1,4-dioxane biodegradation in bioreactors, biofiltration systems, and biosparging systems had demonstrated the feasibility of 1,4-dioxane bioremediation. While challenges remained for both physicochemical and biological processes regarding meeting the stringent drinking water guidelines while operating at relatively low costs. The identification of limiting growth cofactors as well as the pretreatment of chlorinated solvents may be needed to achieve an effective 1,4-dioxane biodegradation to below the drinking water guidelines. Also, treatment chains that combine physical-chemical processes and biodegradation can be considered as a synergistic approach for effective 1,4-dioxane removal as well as reducing the operating cost.