Absolute reduction of perchlorate has proven complex owing to the diverse characteristics of the perchlorate ion. Technologies such as chemical reduction, ozone/peroxide, nanofiltration, and reverse osmosis have had limited success, high costs and are not environmentally friendly. A great deal of research and reviews on ion exchange and biodegradation have been carried out, but conditions for optimal biodegradation are not yet well understood. The acceptability of biological treatment of perchlorate has been limited due to challenges such as electron donor availability, which impacts on the environmental sustainability of perchlorate biodegradation, the biomass inventory, secondary contamination of treated water due to contact with micro-organisms between the treatment unit and the final effluent, and the presence of other energetic compounds such as Royal Demolition Explosives and 2,4-dinitroanisole (DNAN) in army PAX 21 production water and other competing electron acceptors such as nitrate and sulfate. Therefore, the current research concern is about optimization of the biodegradation of perchlorate for large-scale applicability. In addition to summarizing the strengths and weaknesses of developed and emerging perchlorate treatment technologies, this review focuses on research developments in biological treatment of ammonium perchlorate, perchlorate reducing bacteria, factors affecting biodegradation of NH4 ClO4 and previous research recommendations on efficient, effective, and stable biological treatment of perchlorate-contaminated wastewater.

INTRODUCTION

Perchlorate is among the most recently acknowledged group of toxins called endocrine disrupters. It is manufactured as perchloric acid and salts such as ammonium perchlorate, magnesium perchlorate, sodium perchlorate, and potassium perchlorate for use in various processes. These have been significantly introduced into the environment in the form of disinfectants, fertilizers, bleaching agents, blasting agents, herbicides, and rocket propellants. The presence of perchlorate in the environment is detrimental to human health as it can alter hormonal balances and impede human reproduction and development.

Exposure to ammonium perchlorate has been reported to lead to hypothyroidism (Park et al. 2006; Chen et al. 2008). The perchlorate ion competitively inhibits iodide uptake by the thyroid glands, thus affecting the functionality of the thyroid. Naturally occurring chemicals, such as thiocyanate (in food and cigarette smoke) and nitrate (in some food), are also known to inhibit iodide uptake. This potentially leads to a reduction in the production of thyroid hormones; triiodothyronine and thyroxin thyroid hormones (serum T3 and T4) and subsequently, increased production of thyroid stimulating hormones (serum TSH) associated with lower thyroidal hormone storage (Mattie et al. 2006). Thyroid hormones play an important role in regulating metabolism and oxygen consumption. They are critical for normal growth and development in fetuses, infants, and young children.

Short-term exposure to high doses may cause eye and skin irritation, coughing, nausea, vomiting, and diarrhea. Long-term exposure could result in neuro-developmental defects due to decreased thyroid hormones, thyroid hyperplasia resulting from severe/sustained iodine deficiency, and ultimately, growth of tumors due to increased levels of TSH, although tumor growth is a rare occurrence unless in individuals susceptible to disruption of thyroid functions. Research shows that chlorate and chlorite produced during reduction of perchlorate can also cause hemolytic anemia in laboratory animals, methemoglobin formation in mammals, and toxicity in micro-organisms and plants (University of Nebraska 2003; Baldridge et al. 2004; Braverman et al. 2004; National Research Council 2005; Mattie et al. 2006; Park et al. 2006; Chen et al. 2008).

According to the National Research Council, the main adverse effects of perchlorate ingestion are hormonal imbalances, metabolic sequelae such as decreased metabolic rate and slowing of the function of many organ systems at any age, and abnormal fetus and child development. The model in Figure 1 shows the exposure–dose response continuum considered in the context of biomarkers (classified as measures of exposure, effect, and susceptibility) and level of organization at which toxicity is observed (United States Environmental Protection Agency (US EPA) 2008).

Figure 1

Exposure–dose response continuum.

Figure 1

Exposure–dose response continuum.

It should therefore be noted that even at microgram levels, perchlorate causes toxicity to flora and fauna and affects growth, metabolism, and reproduction in humans and animals.

Perchlorate contamination occurs in groundwater, surface water, soil, and drinking water sources. It has also been found in many nutritionally important foods such as dairy and human milk, lettuce and other leafy green vegetables, forage, cereals, cantaloupe, citrus fruits, canned foods, wines, beer, bottled water, and human urine, posing serious public health concerns.

Recent studies on perchlorate pollution status in China indicate widespread contamination. The concentration of perchlorate in sewage sludge, rice, bottled drinking water, and milk in various areas in China was found to be in the range of 0.56–379.9 μg/kg, 0.16–4.88 μg/kg, 0.037–2.013 μg/L, and 0.30–9.1 μg/L, respectively (Shi et al. 2007). In the USA, federal and state agencies identified more than 400 sites including California, Utah, Nevada, Texas, Arkansas, Maryland, Massachusetts, Arizona, New York, the districts of Columbia, and two commonwealths where drinking water, surface water, groundwater, and soil were found to be contaminated with perchlorate, affecting drinking water supplies to more than 20 million people (US EPA 2008). Perchlorate contamination has also been found in other parts of the world such as in Israel, where widespread perchlorate contamination has been found in the 40 m deep vadose zone near an ammonium perchlorate manufacturing plant north of Tel Aviv, above the central part of Israel's coastal aquifer, with peak concentrations of 1,200 mg kg−1 sediment (Gal et al. 2008).

Ammonium perchlorate [AP (NH4ClO4)] is a form of high grade perchlorate domestically produced for use in the Department of Defense (DoD), and the National Aeronautics and Space Administration (NASA). It has been used by the US DoD as an oxidizer in munitions and missiles since the 1940s. Ammonium perchlorate has a limited shelf life, so inventories must be periodically replaced with a fresh supply, creating large quantities of ammonium perchlorate waste that need to be disposed of. For instance, the end of the Cold War left the US DoD with a projected 140 million pounds of rocket propellant to be disposed of between 1993 and 2005 (US EPA 2009).

The ammonium ions initially present in the ground at contaminated sites generally biodegrade over time, whereas the perchlorate ion persists due to its poor reactivity and great solubility and mobility in water. Furthermore, perchlorate compounds and the perchlorate anion do not volatilize from water to air. As a result, perchlorate plumes in groundwater can be extensive. For example, the perchlorate plume at a former safety flare site (the Olin Flare Facility) in Morgan Hill, California, extends more than nine miles (US EPA 2009).

PERCHLORATE OCCURRENCE AND SOURCES OF CONTAMINATION

Natural occurrence

Perchlorate is both a naturally occurring and man-made anion . It occurs naturally, especially in arid regions, and can be found as a natural impurity in nitrate salts used to manufacture nitrate fertilizers. Contamination from military and industrial sources is well documented but up to now natural background levels are not well understood. A number of possible processes through which perchlorate can be naturally formed are being studied. Figure 2 shows some of these processes and possible pathways of perchlorate formation. Highly oxidizing species such as ClO and can be generated from stable Cl minerals by the action of UV and from ClO4 minerals by the action of cosmic γ and x-rays (Kounaves et al. 2013).

Figure 2

Processes and pathways for the production of perchlorate and accompanying intermediary oxychlorines (EPSC Abstracts Vol. 8, EPSC2013-799-1 2013).

Figure 2

Processes and pathways for the production of perchlorate and accompanying intermediary oxychlorines (EPSC Abstracts Vol. 8, EPSC2013-799-1 2013).

Currently, atmospheric formation by photochemical reactions between chloride and ozone and perhaps ultraviolet radiation is the working theory. After atmospheric formation, perchlorate is dissolved in precipitation and returns to the Earth's surface. Atmospheric deposition occurs from rain, washing compounds out of the atmosphere, and the settling out of dry airborne materials, including dust.

In arid environments, where the rate of deposition exceeds the rate of dissolution by ongoing precipitation, perchlorate can be incorporated into certain geologic formations (Orris 2002), as shown in Figure 3. The samples deriving from soil-caliche show 100% with perchlorate, and the proportion of halite samples is 44%.

Figure 3

Comparative assessment of the environmental sustainability of perchlorate treatment technologies for drinking water using consumables as the driving force (Choe et al. 2013).

Figure 3

Comparative assessment of the environmental sustainability of perchlorate treatment technologies for drinking water using consumables as the driving force (Choe et al. 2013).

The rate of atmospheric deposition has long been thought to be small, however recent studies show that perchlorate atmospheric deposition rates are about 10 times larger than previously reported (Andraski et al. 2014). In this study, atmospheric deposition flux was 343 mg ha−1 yr−1, which is approximately 10 times that for published southwestern wet-deposition fluxes. The study further proved that both wet and dry atmospheric depositions are important contributors of perchlorate to the land surface.

The distribution of this naturally occurring in deserts, arid and semi-arid regions has mainly been related to evaporative concentration and unsaturated transport. This process leads to higher and other ion concentrations in groundwater where the water table is relatively shallow, and in areas with lower saturated thickness (Böhlke et al. 2005; Balaji et al. 2007; Jackson et al. 2010; Andraski et al. 2014).

Natural perchlorate deposits

Perchlorate has been known to be present naturally in nitrate deposits in the Atacama Desert of northern Chile since the 1880s. Initially, it was believed that the extensively documented Chilean deposits were the only source of naturally occurring perchlorate. However, based on isotopic composition, natural perchlorate indigenous to the United States and other parts of the world can be distinguished from both synthetic perchlorate and perchlorate derived from Chilean fertilizers (Böhlke et al. 2005; Jackson et al. 2010). Various researches recently carried out show widespread occurrence of natural perchlorate independent of the Chilean deposits. These are summarized in Table 1 and serve as confirmation that perchlorate is produced globally and continuously in the Earth's atmosphere, that it typically accumulates in hyper arid areas, and that it does not build up in oceans or other wet environments probably due to microbial reduction.

Table 1

Natural perchlorate occurrence

Site Analyzed substances Perchlorate concentrations Source of perchlorate Reference 
Amargosa Desert, Nevada Soil, leaves from shrubs, rain, dust Shallow soils had high levels of perchlorate – 10–20 grams per hectare in the top 30 cm Atmospheric wet and dry deposition Andraski et al. (2014)  
Southwestern United States (arid and semi-arid) Discrete depth subsurface soil and sediment Peak concentrations ranged from 1.6 to 13 micrograms per kilogram dry solid (μg kg−1Atmospherically deposited chloride (ClBalaji et al. (2007); Jackson et al. (2010)  
High plains of Texas and New Mexico Groundwater Generally low (<4 ppb), although some areas are impacted by concentrations up to 200 ppb Atmospheric deposition and meteoric ClO4 that accumulated in the unsaturated zone over the last 2–10,000 years Rajagopalan et al. (2006)  
Antarctic Soil and ice from several Antarctic dry valleys Concentrations reach up to 1,100 μg/kg Atmospheric deposition Kounaves et al. (2010)  
North America (Continental United States, Alaska, and Puerto Rico) 1,578 composite wet deposition samples Perchlorate concentrations varied from <5 ng/L to a high of 102 ng/L with a mean of 14.1 ± 13.5 ng/L Wet atmospheric deposition Rajagopalan et al. (2009)  
  Annual perchlorate flux averaged at 65 ± 30 mg/ha-year and ranging from 12.5 (TX) to 157 mg/ha-year   
Arctic Devon Island ice cap Concentrations varied between 1 and 18 ng L−1 Stratospheric formation by chlorine radicals reacting with ozone year round (perchlorate concentrations the total ozone level) Furdui & Tomassini (2010)  
   Tropospheric formation (perchlorate was correlated with the chloride concentrations during summer)  
Chilean Atacama Desert Cultivated and uncultivated soils, superficial running water Concentrations ranging from 290 ± 1 to 2,565 ± 2 μg/kg in soils with highest concentration in humber stone soil. Concentrations in water ranged from 744 ± 0.01 to 1,480 ± 0.02 μμg/L Wet and dry deposition Calderón et al. (2014)  
Site Analyzed substances Perchlorate concentrations Source of perchlorate Reference 
Amargosa Desert, Nevada Soil, leaves from shrubs, rain, dust Shallow soils had high levels of perchlorate – 10–20 grams per hectare in the top 30 cm Atmospheric wet and dry deposition Andraski et al. (2014)  
Southwestern United States (arid and semi-arid) Discrete depth subsurface soil and sediment Peak concentrations ranged from 1.6 to 13 micrograms per kilogram dry solid (μg kg−1Atmospherically deposited chloride (ClBalaji et al. (2007); Jackson et al. (2010)  
High plains of Texas and New Mexico Groundwater Generally low (<4 ppb), although some areas are impacted by concentrations up to 200 ppb Atmospheric deposition and meteoric ClO4 that accumulated in the unsaturated zone over the last 2–10,000 years Rajagopalan et al. (2006)  
Antarctic Soil and ice from several Antarctic dry valleys Concentrations reach up to 1,100 μg/kg Atmospheric deposition Kounaves et al. (2010)  
North America (Continental United States, Alaska, and Puerto Rico) 1,578 composite wet deposition samples Perchlorate concentrations varied from <5 ng/L to a high of 102 ng/L with a mean of 14.1 ± 13.5 ng/L Wet atmospheric deposition Rajagopalan et al. (2009)  
  Annual perchlorate flux averaged at 65 ± 30 mg/ha-year and ranging from 12.5 (TX) to 157 mg/ha-year   
Arctic Devon Island ice cap Concentrations varied between 1 and 18 ng L−1 Stratospheric formation by chlorine radicals reacting with ozone year round (perchlorate concentrations the total ozone level) Furdui & Tomassini (2010)  
   Tropospheric formation (perchlorate was correlated with the chloride concentrations during summer)  
Chilean Atacama Desert Cultivated and uncultivated soils, superficial running water Concentrations ranging from 290 ± 1 to 2,565 ± 2 μg/kg in soils with highest concentration in humber stone soil. Concentrations in water ranged from 744 ± 0.01 to 1,480 ± 0.02 μμg/L Wet and dry deposition Calderón et al. (2014)  

Anthropogenic sources

Perchlorate is used in a variety of commercial, chemical, and industrial processes. The pervasive contamination of perchlorate arises from its application in a wide range of processes. Salts of chlorate have been used as defoliants, leading to speculation that these could be sources of perchlorate in groundwater (Brown & Gu 2005).

Ammonium perchlorate has been used as a primary oxidizer in solid propellants for missiles, munitions, and rockets for many years now because it has a high oxygen content and decomposes to the gaseous phase's products, water, HCI, N2, and 02, leaving no residue. The one disadvantage in its use as an oxidizing agent is that it does not function well in solid fueled rockets after it adsorbs too much water, requiring significant proper disposal and replenishment (Brown & Gu 2005). When rockets are successfully launched, the intense heat leads to nearly complete reaction of the perchlorate. Therefore, release of perchlorate to the environment often occurs when its intended use does not occur, such as in dismantling and disposal of rockets, accidental release from manufacturing facilities, or unsuccessful rocket launches.

Currently, no suitable replacements for perchlorate as an oxidizer have been found that can satisfy defense/military, aeronautics and space administration logistics, performance, safety and environmental requirements. Studies on the use of ammonium dinitramide and other nitramines and nitramides are ongoing, although existing alternative energetic oxidizers show significant cost, availability, and performance issues if used in fielded weapon systems (Strategic Environmental Research and Development Program 2004a, 2004b). Owing to their use in military applications, many countries consider the amounts that they make confidential.

Overall, perchlorate salts are essential constituents of composite propellants, underwater explosives, and pyrotechnic compositions. These are used in fireworks, airbag initiators for vehicles, matches, signal flares, blasting agents, and in some disinfectants. Perchlorate contamination investigation in groundwater and surface water from Sivakasi and Madurai in the Tamil Nadu State of South India showed that concentrations of perchlorate were <0.005–7,690 μg/L in groundwater, <0.005–30.2 μg/L in surface water, and 0.063–0.393 μg/L in tap water. Levels in groundwater were significantly higher in the fireworks factory area than in the other locations, indicating that fireworks and safety match industries are principal sources of perchlorate pollution (Isobe et al. 2013). In agriculture they are used in fertilizers, as additives in cattle feed, and have also been found in some herbicides as an incidental byproduct in the manufacture of sodium chlorate used in agricultural herbicides and defoliants. In textile industries, perchlorate has been found in finished leather, fabric fixers, and dyes. Further, perchlorate salts are used in lubricating oils, electroplating, aluminum refining, the manufacture of rubber, paint, enamel production, and magnesium batteries (Motzer 2001). Pulp and paper industries which use sodium chlorate are also possible sources of perchlorate contamination. For use in the paper industry, sodium chlorate is converted to C102 by reduction with hydrogen peroxide in the presence of sulfuric acid. The C102 is then used as a bleaching agent.

It has been reported that laboratory grade sodium chlorate contains 0.2% perchlorate on a weight basis, while the analytical reagent grade chemical has 200–900 ppm (Burns et al. 1989).

Although authoritative reference works indicate that technical grade sodium chlorate is 99.5% pure (McKetta & Weismantel 1995), given the large quantities of sodium chlorate used annually by the pulp and paper industry, sodium chlorate cannot be ignored as a possible source of perchlorate contamination in the environment.

In farming regions, the main source of perchlorate contamination of soil is fertilizers. However, very high levels of perchlorate contamination of soil and sediments have been reported to occur from DoD and other federal facilities. In groundwater contamination, perchlorate occurs as a result of percolation of rainwater through contaminated sand or soil. It also may be released to surface water from runoff or erosion of contaminated sand or soil. Furthermore, water-gel and emulsion blasting agents containing perchlorates used in difficult rock blasting, underground and trenching, deep boreholes, and other applications that require extra energy over conventional agents, could lead to a higher incidence of incomplete combustion and become a source of perchlorate contamination if applied where surface and groundwater can be affected.

Human exposure

Ingestion of contaminated food, milk, and drinking water has been reported to be the primary pathway for human exposure to perchlorate. Other modes of exposure, although minor, include adsorption through the skin and inhalation. Perchlorate being an inorganic compound and completely ionized in water, the potential for dermal absorption through intact skin is unlikely.

However, the primary pathway for workers in industrial and commercial production facilities or commercial use of perchlorate salts is inhalation of ammonium perchlorate dust. Occupational exposure in ammonium perchlorate production facilities has been shown to be higher than potential exposures from drinking water or food sources (Gibbs et al. 1998; Braverman et al. 2005).

PERCHLORATE DRINKING WATER STANDARDS

The amount of perchlorate safe for humans is an issue of major scientific deliberation. In addition to perchlorate contamination being detected in the United States, China, and Israel, it has also been detected in Japan's Tome river watershed (Kosaka et al. 2007) in concentrations ranging from 0.08 to 2,300 mg/L in the upper watershed and 0.73–25 mg/L in the middle and lower portions of the watershed and in Korea (Quinones et al. 2007) at concentrations ranging from 0.15 to 60 mg/L in the Nakdong river watershed and 0.08–2.3 mg/L in the Yongsan river. Despite perchlorate contamination being an international problem, most nations lack definitive drinking water regulations for perchlorate.

Following the discovery of perchlorate in drinking water sources in various states in the USA and recognition that a perchlorate dose of 6 mg kg−1 body weight d−1 or more administered to hyperthyroidism patients over a 2-month period could lead to fatal bone marrow disorders in 1992, perchlorate was added to the EPA contaminant candidate list in 1998 (US EPA 2009).

After a number of studies, researches, and consultations among the National Research Council, DoD, National Academy of Science, and the Environmental Protection Agency in January 2006, the EPA Superfund office issued guidance for a drinking water-equivalent level of 24.5 μg/L perchlorate to be considered as the preliminary remediation goal (PRG) in order to minimize health risks (Kucharzyk et al. 2009). The EPA's Children's Health Protection Advisory Committee (CHPAC) however suggested that the PRG of 24.5 μg/L does not protect infants, who are highly susceptible to neuro-developmental toxicity and may be more exposed than fetuses to perchlorate. While noting that perchlorate is concentrated in breast milk and that nursing infants could receive daily doses greater than the reference dose if the mother was exposed to 24.5 ppb perchlorate in tap water, the CHPAC recommended the level to be lowered.

In October 2008 the EPA issued an Interim Drinking Water Health Advisory for perchlorate of 15 ppb and announced a delay on the decision about whether to set a drinking water standard for perchlorate until it received advice from the NRC (US EPA 2009).

As of 2009, no national drinking water standard for perchlorate had been set. The promulgation of a national primary drinking water regulation for perchlorate has hence proven to be a multi-year process, especially due to economic considerations.

Owing to the EPA's reluctance to set a national drinking water standard for perchlorate, individual states have set their own advisory levels, as shown in Table 2.

Table 2

Drinking water standards for perchlorate

State Drinking water perchlorate standard 
Nevada 18 μg/L 
Arizona 14 μg/L 
New York 5 μg/L 
California 6 μg/L 
Texas 4 μg/L 
New Mexico 1 μg/L 
Massachusetts 1 μg/L 
Maryland 1 μg/L 
State Drinking water perchlorate standard 
Nevada 18 μg/L 
Arizona 14 μg/L 
New York 5 μg/L 
California 6 μg/L 
Texas 4 μg/L 
New Mexico 1 μg/L 
Massachusetts 1 μg/L 
Maryland 1 μg/L 

TREATMENT METHODS

Wastewater and drinking water treatment to remove/destroy perchlorate can be achieved by both biotic and abiotic methods. Perchlorate is highly toxic, soluble, relatively stable, and mobile in water. It is also highly soluble in polar organic solvents, slow to react, and ammonium perchlorate has a limited shelf life. Table 3 summarizes the physical and chemical properties of common perchlorate salts.

Table 3

Physical and chemical properties of perchlorate compounds

Property Ammonium perchlorate Sodium perchlorate Potassium perchlorate Perchloric acid 
CAS No. 7790-98-9 7601-89-0 7778-74-7 7601-90-3 
Formula NH4ClO4 NaClO4 KClO4 HClO4 
Formula weight 117.49 122.44 138.55 100.47 
Color/Form White, orthorhombic crystals White, orthorhombic crystals; white deliquescent crystals Colorless crystals or white, crystalline powder; colorless, orthorhombic crystals Colorless, oily liquid 
Melting point Decomposes/explodes 480 °C 525 °C −112 °C 
Density 1.95 g/cm3 2.52 g/cm3 2.53 g/cm3 1.768 g/cm3 
Solubility 200 g/L of water at 25 °C 209.6 g/100 mL of water at 25 °C 15 g/L of water at 25 °C Miscible in cold water 
Additional solubility information Soluble in methanol; slightly soluble in ethanol, acetone; almost insoluble in ethyl acetate, ether 209 g/100 mL water at 15 °C; 284 g/100 mL water at 50 °C; soluble in alcohol Soluble in 65 parts cold water, 15 parts boiling water; practically insoluble in alcohol; insoluble in ether Not provided 
Property Ammonium perchlorate Sodium perchlorate Potassium perchlorate Perchloric acid 
CAS No. 7790-98-9 7601-89-0 7778-74-7 7601-90-3 
Formula NH4ClO4 NaClO4 KClO4 HClO4 
Formula weight 117.49 122.44 138.55 100.47 
Color/Form White, orthorhombic crystals White, orthorhombic crystals; white deliquescent crystals Colorless crystals or white, crystalline powder; colorless, orthorhombic crystals Colorless, oily liquid 
Melting point Decomposes/explodes 480 °C 525 °C −112 °C 
Density 1.95 g/cm3 2.52 g/cm3 2.53 g/cm3 1.768 g/cm3 
Solubility 200 g/L of water at 25 °C 209.6 g/100 mL of water at 25 °C 15 g/L of water at 25 °C Miscible in cold water 
Additional solubility information Soluble in methanol; slightly soluble in ethanol, acetone; almost insoluble in ethyl acetate, ether 209 g/100 mL water at 15 °C; 284 g/100 mL water at 50 °C; soluble in alcohol Soluble in 65 parts cold water, 15 parts boiling water; practically insoluble in alcohol; insoluble in ether Not provided 

Source: National Library of Medicine. Specialized Information Services. 2004. Hazardous Substances Data Bank.

Perchlorate compounds and the perchlorate anion do not volatilize from water to air. Owing to these properties, most well-known processes have had little success especially with respect to environmental and economic considerations.

Previous reviews on most physical and chemical processes, namely, chemical and electrochemical reduction, adsorption on activated carbon, granular activated carbon (GAC) or metal ions, ion exchange, membrane processes such as reverse osmosis (RO), electrodialysis, ultrafiltration (UF) and nanofiltration (NF) indicate that these have had limited application in perchlorate removal. This is attributed to high costs, non-selectivity, slow reduction rates, rapid accumulation of active sites in the presence of competing anions like nitrates, their need for extreme reaction conditions (temperature, pressure, and pH), and generation of large volumes of concentrated waste streams (with perchlorate and total dissolved solids (TDS)) that cannot be easily disposed of or require further reduction processes before disposal (Hatzinger 2005; Bardiya & Bae 2011; Ye et al. 2012). For instance, electrochemical reduction, despite showing great promise in complete destruction of the perchlorate ion, is inapplicable on a large scale as it is very slow, as is chemical reduction. Ion exchange merely transfers perchlorate from water to the resin, making it a non-selective and incomplete process since it is a physico-chemical process based on exchanging an anion (typically Cl) with the perchlorate ion in the water as shown by the equation below: 
formula
In a review, Bardiya & Bae (2011) found that perchlorate-laden spent resins with perchlorate 200–500 mg L−1 required regeneration resulting in the production of concentrated brine 6–12% NaCl or caustic waste streams. Further research and subsequent application of selective anion exchange is fortunately progressing well.

Evidence obtained from a growing number of bench-scale tests also show the potential effectiveness of phytoremediation of perchlorate-contaminated soils, surface, and groundwater (Susarla et al. 1999a, 1999b, 1999c, 2000; Nzengung et al. 2001, 2004; Nzengung & Wang 2000; Nzengung & McCutcheon 2003).

Laboratory research involving testing with several wetland species, including Typha latifolia (cattail), Spirodela polyrhiza (L.) Shield (duck weed), microbial mats, and Myriophyllum aquaticum (parrot feather), as well as several terrestrial plants, including black willow (Salix nigra and Salix caroliniana), eastern cottonwood (Populus deltoides), eucalyptus (Eucalyptus cinerea), loblolly pine (Pinus taeda), French tarragon (Artemisia dracunculus), and spinach (Spinacia oleracea) (Susarla et al. 1999a, 1999b, 1999c, 2000; Nzengung & Wang 2000; Nzengung et al. 2001, 2004; Nzengung & McCutcheon 2003) have been carried out.

Bench-scale studies have identified the predominant mechanisms of phytoremediation of perchlorate as: (1) uptake and phytodegradation, (2) uptake and phytoaccumulation by some plant species, and (3) rapid rhizodegradation. As uptake and phytodegradation is a slower process, it poses an ecological risk resulting from the temporal phytoaccumulation of some fraction of the perchlorate being taken up and transported mainly to plant leaves. In bench-scale tests, uptake may account for the removal of 5–25% of the perchlorate present in the root zone of plants (Susarla et al. 2000).

Constructed wetlands are also increasingly being used for the remediation of groundwater or surface water impacted by industrial chemicals and wastes such as landfill leachate and explosives such as TNT or Royal Demolition Explosives (RDX). The increased application is due to the low capital and operation and maintenance (O&M) costs associated with this mostly passive technology. Recent successful trials of small-scale wetland reactors suggest that full-scale use of constructed wetlands could be a cost-effective method to deal with large volumes of perchlorate-contaminated water sources such as groundwater.

Membrane technologies which employ a semi-permeable membrane that prevents the passage of certain ions to treat water, such as UF, NF, and RO, although having been reported to be effective for perchlorate removal are not suitable for large-scale applications because of fouling issues, costliness, and the generation of large volumes of reject streams. Electrodialysis, being the most effective membrane technology, has extremely high operational costs (Srinivasan & Sorial 2009). Table 4 summarizes the strengths and weaknesses of various perchlorate treatment technologies.

Table 4

Strengths and weaknesses of perchlorate treatment technologies

Technology Strengths Weaknesses Other remarks 
Tailored granular activated carbon (T-GAC) Proven effective technology Not a destructive technology Spent carbon (all adsorptive sites used) can be disposed of by landfill or incineration. Landfill may cause perchlorate to desorb from the carbon and contaminate off-site areas whereas incineration destroys the perchlorate ion and reduces the GAC to a small amount of ash with no secondary toxic contaminants to manage 
 No regeneration brine is created during treatment – a highly adsorbent carbon material activated by heating and coated with a thin layer of a surface-active substance is used to adsorb the negatively charged perchlorate ion Carbon tailoring has limited adsorption capacity for perchlorate removal hence not effective for large-scale application and high perchlorate concentrations  
 Can be regenerated for reuse Quaternary ammonium monomers improve capacity over what is achieved with organic polymers but have not yet been approved for use in potable water treatment  
  Not as effective as other technologies, hence is only best as retrofit of existing systems such as IX  
  Spent GAC may require treatment prior to ordinary or hazardous waste disposal  
  GAC adsorption for perchlorate might require pretreatment for removal of suspended solids, silica, or mica from streams to be treated  
  Ability to remove perchlorate can be reduced by water-soluble co-contaminants with a high polarity  
Ion exchange (IX) regenerate systems Proven technology for large-scale application that can be used for both high and low CLO4 concentrations as well as groundwater with high total solids (TS), though is more cost-effective when the perchlorate concentration is low Costly non-destructive technology Can be used as a polishing step for biological treatment processes or electrodialysis 
 Regenerable systems offer the advantage of small footprints, high regeneration efficiencies, and automated operation Non-selectivity as it generates perchlorate-laden brine waste streams that require additional treatment, hence can be highly costly  
  Resin beds can be clogged by organics, TDS, calcium, or iron in the influent  
Ion exchange single use systems (use of non-regenerable resins) Proven technology for large-scale application that can be used for both high and low CLO4 concentrations as well as groundwater with high TS Costly and non-destructive technology Selectivity (use of resins with high affinities for perchlorate) would reduce long-term operating costs 
 Do not generate a perchlorate-laden waste stream (brine) that is created during resin regeneration Need substitution of exhausted resins, which must be removed from the facility and sent for disposal  
  Resins can cause fouling, plugging, channeling, bacterial contamination, agglomeration, and compaction problems  
  Resin beds can be clogged by organics, TDS, calcium, or iron in the influent  
Electrochemical – capacitive deionization More energy-efficient than competing technologies like thermal processes Non-destructive, very slow and highly costly, hence not applicable for large scale The resulting brine can be treated through catalytic treatment and biological reduction 
 Uses less energy with lower operating pressures than RO Small electrochemical energy limit capacity, hence is efficient only on low perchlorate concentrations  
 No bothersome membranes Regeneration of the electrodes yields concentrated brine  
Phytoremediation (use of plants to remove contaminants from soil and groundwater) Low cost high public acceptance Relatively slow process Choice of the plant species is important and there is limited research and data 
 Little disturbance to the environment Possible ecological risk due to temporal phytoaccumulation of some fraction of the perchlorate taken up and transported mainly to plant leaves and finally/also possibility of inclusion into the food chain  
 Can treat other common co-contaminants, such as chlorinated solvents, explosives Depth and climatic restrictions as climate greatly impacts plant growth  
 No secondary waste production   
Electrolysis Destruction of perchlorate into harmless by-products and leaves the water contaminant free High electrical energy requirements Still an emerging technology, not fully proven yet; limited research 
 No brine or other waste stream   
Ultraviolet laser reduction Complete destruction of low perchlorate concentrations Not effective for high perchlorate concentrations Still an emerging technology, not fully proven yet; limited research 
 Catalytic reduction is faster than biological reduction   
Membrane technologies   
RO (<0.0001 microns pore sized membrane) Proven and effective in perchlorate removal Non-destructive technology hence requires further treatment of reject To ensure water palatability post-treatment may require sodium chloride or sodium bicarbonate 
 Can treat high-TDS water and concentrated brines High capital, O&M costs  
 Automated systems Only applicable as standalone for low perchlorate concentrations  
 Can be used as a pretreatment or polishing technology for other systems Generate high volumes of reject hence unsuitable for large scale  
  Non-ionic selectivity in the semi-permeable membrane can alter the pH of the effluent stream and make it corrosive  
  Membrane resilience and fouling  
  Needs higher operating pressure than other membrane technologies due to small pore sizes of membrane filters, hence increased need for power supply  
Ultrafiltration (0.1–0.005 microns pore size); nanofiltration (0.005–0.0001 microns pore size) Lower energy requirements than RO due to large pore sizes UF membrane pore sizes are too large to remove perchlorate (0.00035 microns)  
  Nanofiltration membrane has limited success in perchlorate ion removal  
  Expensive, fouling issues and waste streams that require further management  
Electrodialysis Most effective membrane technology Non-destructive and generates contaminated waste streams that require further treatment/disposal When high TDS is a consideration, pretreatment with IX resin membranes is possible 
 Can manage high TDS Extremely high operation/energy costs  
 Capable of high recovery (more product and less brine than IX) and is not affected by non-ionic substances such as silica The resulting concentrate may require larger volumes of water for treatment before disposal  
  Permeable membrane has low selectivity for perchlorate ions  
  Fouling issues  
Technology Strengths Weaknesses Other remarks 
Tailored granular activated carbon (T-GAC) Proven effective technology Not a destructive technology Spent carbon (all adsorptive sites used) can be disposed of by landfill or incineration. Landfill may cause perchlorate to desorb from the carbon and contaminate off-site areas whereas incineration destroys the perchlorate ion and reduces the GAC to a small amount of ash with no secondary toxic contaminants to manage 
 No regeneration brine is created during treatment – a highly adsorbent carbon material activated by heating and coated with a thin layer of a surface-active substance is used to adsorb the negatively charged perchlorate ion Carbon tailoring has limited adsorption capacity for perchlorate removal hence not effective for large-scale application and high perchlorate concentrations  
 Can be regenerated for reuse Quaternary ammonium monomers improve capacity over what is achieved with organic polymers but have not yet been approved for use in potable water treatment  
  Not as effective as other technologies, hence is only best as retrofit of existing systems such as IX  
  Spent GAC may require treatment prior to ordinary or hazardous waste disposal  
  GAC adsorption for perchlorate might require pretreatment for removal of suspended solids, silica, or mica from streams to be treated  
  Ability to remove perchlorate can be reduced by water-soluble co-contaminants with a high polarity  
Ion exchange (IX) regenerate systems Proven technology for large-scale application that can be used for both high and low CLO4 concentrations as well as groundwater with high total solids (TS), though is more cost-effective when the perchlorate concentration is low Costly non-destructive technology Can be used as a polishing step for biological treatment processes or electrodialysis 
 Regenerable systems offer the advantage of small footprints, high regeneration efficiencies, and automated operation Non-selectivity as it generates perchlorate-laden brine waste streams that require additional treatment, hence can be highly costly  
  Resin beds can be clogged by organics, TDS, calcium, or iron in the influent  
Ion exchange single use systems (use of non-regenerable resins) Proven technology for large-scale application that can be used for both high and low CLO4 concentrations as well as groundwater with high TS Costly and non-destructive technology Selectivity (use of resins with high affinities for perchlorate) would reduce long-term operating costs 
 Do not generate a perchlorate-laden waste stream (brine) that is created during resin regeneration Need substitution of exhausted resins, which must be removed from the facility and sent for disposal  
  Resins can cause fouling, plugging, channeling, bacterial contamination, agglomeration, and compaction problems  
  Resin beds can be clogged by organics, TDS, calcium, or iron in the influent  
Electrochemical – capacitive deionization More energy-efficient than competing technologies like thermal processes Non-destructive, very slow and highly costly, hence not applicable for large scale The resulting brine can be treated through catalytic treatment and biological reduction 
 Uses less energy with lower operating pressures than RO Small electrochemical energy limit capacity, hence is efficient only on low perchlorate concentrations  
 No bothersome membranes Regeneration of the electrodes yields concentrated brine  
Phytoremediation (use of plants to remove contaminants from soil and groundwater) Low cost high public acceptance Relatively slow process Choice of the plant species is important and there is limited research and data 
 Little disturbance to the environment Possible ecological risk due to temporal phytoaccumulation of some fraction of the perchlorate taken up and transported mainly to plant leaves and finally/also possibility of inclusion into the food chain  
 Can treat other common co-contaminants, such as chlorinated solvents, explosives Depth and climatic restrictions as climate greatly impacts plant growth  
 No secondary waste production   
Electrolysis Destruction of perchlorate into harmless by-products and leaves the water contaminant free High electrical energy requirements Still an emerging technology, not fully proven yet; limited research 
 No brine or other waste stream   
Ultraviolet laser reduction Complete destruction of low perchlorate concentrations Not effective for high perchlorate concentrations Still an emerging technology, not fully proven yet; limited research 
 Catalytic reduction is faster than biological reduction   
Membrane technologies   
RO (<0.0001 microns pore sized membrane) Proven and effective in perchlorate removal Non-destructive technology hence requires further treatment of reject To ensure water palatability post-treatment may require sodium chloride or sodium bicarbonate 
 Can treat high-TDS water and concentrated brines High capital, O&M costs  
 Automated systems Only applicable as standalone for low perchlorate concentrations  
 Can be used as a pretreatment or polishing technology for other systems Generate high volumes of reject hence unsuitable for large scale  
  Non-ionic selectivity in the semi-permeable membrane can alter the pH of the effluent stream and make it corrosive  
  Membrane resilience and fouling  
  Needs higher operating pressure than other membrane technologies due to small pore sizes of membrane filters, hence increased need for power supply  
Ultrafiltration (0.1–0.005 microns pore size); nanofiltration (0.005–0.0001 microns pore size) Lower energy requirements than RO due to large pore sizes UF membrane pore sizes are too large to remove perchlorate (0.00035 microns)  
  Nanofiltration membrane has limited success in perchlorate ion removal  
  Expensive, fouling issues and waste streams that require further management  
Electrodialysis Most effective membrane technology Non-destructive and generates contaminated waste streams that require further treatment/disposal When high TDS is a consideration, pretreatment with IX resin membranes is possible 
 Can manage high TDS Extremely high operation/energy costs  
 Capable of high recovery (more product and less brine than IX) and is not affected by non-ionic substances such as silica The resulting concentrate may require larger volumes of water for treatment before disposal  
  Permeable membrane has low selectivity for perchlorate ions  
  Fouling issues  

Generally, technologies that have been successfully used to treat perchlorate contamination have primarily involved anion exchange or biological treatment, although anion exchange requires supplementary treatment before disposal. This need for auxiliary treatment of the concentrated brines has resulted in great interest in chemical reduction for many seeking economical and efficient ways of remediation of perchlorate contamination in water. Chemical reduction has been limited due to the fact that perchlorate does not appreciably react at ambient temperature with chemical reducing agents commonly used in remediation, such as sulfite, dithionite, or zero-valent iron (ZVI). Determining a stable remediation procedure has been challenged by the discovery that some weaker reducing agents react with perchlorate at measurable rates, whereas stronger reducing agents fail to react at all.

Despite biological reduction and ion exchange being more established than other methods, there is still not a single technology that can be directly applied to a drinking water treatment system for complete removal of perchlorate (Srinivasan & Sorial 2009), an indication that an integration of these technologies may have to be adopted for a completely effective and economical perchlorate reduction.

Choe et al. (2013) analyzed the environmental sustainability of ion exchange, biodegradation, and catalytic reduction perchlorate reducing technologies using a life cycle assessment. Resource consumption/consumable during the operation phase was used as the major driving force for environmental impacts. The analysis indicated that the environmental impacts of heterotrophic biological treatment were 2–5 times more sensitive to influent conditions (i.e., nitrate/oxygen concentration) and were 3–14 times higher compared to ion exchange. Catalytic treatment using carbon-supported Re–Pd had a higher (ca. 4,600 times) impact than others, but was within 0.9–30 times the impact of ion exchange with a newly developed ligand-complexed Re–Pd catalyst formulation. Autotrophic biological treatment was the most environmentally beneficial among all methods (Choe et al. 2013).

Figure 3 shows the three perchlorate reducing technologies and their potential impact on the environment measured as kg CO2 equivalent per kg of treated, an indication of the global warming potential of these technologies.

BIODEGRADATION OF AMMONIUM PERCHLORATE

Microbial reduction of perchlorate was observed as early as the 1950s, but research into this has only been conducted during the recent past several years. Biodegradation was recognized by the Air Force Research Laboratory (AFRL) in 1989 as a potential process for treating dilute ammonium perchlorate waste streams and for remediating contaminated soil and groundwater.

Perchlorate can be anaerobically biodegraded under reducing conditions. In these reactions, perchlorate serving as an electron acceptor is readily reduced to water, carbon dioxide, and chloride in the presence of an appropriate food source (electron donor) and redox conditions. The initial steps in perchlorate destruction are the reduction of perchlorate to chlorate and subsequently the reduction of chlorate to chlorite (Figure 4). These steps are mediated by chlorate reductase and are followed by the dismutation of chlorite into chloride and O2, which is catalyzed by a conserved enzyme known as chlorite dismutase (Figure 5). The dismutation of chlorite to chlorine and oxygen is known to be common to all perchlorate-reducing bacteria (PRB). A widely accepted perchlorate-reducing pathway, especially in the use of acetate, a common electron donor, is: 
formula
Figure 4

Wolinella succinogenes HAP 1 metabolic perchlorate-reducing pathway.

Figure 4

Wolinella succinogenes HAP 1 metabolic perchlorate-reducing pathway.

Figure 5

Enzymes involved in biological perchlorate reducing pathway.

Figure 5

Enzymes involved in biological perchlorate reducing pathway.

Both aerobic and anaerobic biological wastewater treatment can be achieved under either suspended growth systems or attached growth systems.

Suspended-growth systems include systems like continuous-flow stirred-tank reactors (CSTRs), most sequencing batch reactors (SBRs), and activated sludge systems (ASSs).

Attached growth systems include biological trickling filters and rotating biological contactors, which were developed to improve the functionality of trickling filter and biological tower systems following the discovery of lightweight plastic media. Owing to the greater heights, trickling filters using plastic media are often termed biological towers. Trickling filter systems have also been described as trickle filters, trickling biofilters, biofilters, biological filters, biological trickling filters, roughing filters, intermittent filters, packed media bed filters for packed bed reactors (PBRs), alternative septic systems, percolating filters, attached growth processes, fluidized bed reactors, and fixed film processes, depending on system characteristics.

The key advantage of attached growth systems over suspended growth reactors is their ability to maintain high densities of biomass within the reactor, even in the presence of rapidly flowing groundwater or wastewater, hence there preferred application in groundwater and drinking water perchlorate treatment. Suspended growth systems have mainly been studied and applied in high strength and industrial perchlorate treatment.

Other biological wastewater treatment systems applicable include the moving bed biological reactor (MBBR), which is an attached growth activated sludge process designed to achieve a high quality effluent (20BOD/30SS/ammonia) within a small footprint with low capital cost, low sludge production, and no return activated sludge stream requirement; and the integrated fixed-film activated sludge system (IFAS), which is an integration of the biofilm carrier technology (MBBR) within a conventional ASS.

Perchlorate degradation to levels needed for drinking water has been achieved using fixed and fluidized bed bioreactors, with most studies conducted in the laboratory. Biological removal of perchlorate has also been evaluated in biofilm reactors using different carrier media including plastic, sand, Celite, and GAC for fixed bed reactors, and sand and GAC for fluidized bed reactors. Advantages of using GAC as a carrier medium include the widespread application of GAC in drinking water treatment plants where existing GAC filters can easily be retrofitted to operate as biologically active carbon (BAC) reactors (Brown et al. 2003; Min et al. 2004; Choi et al. 2007). GAC supports the growth of biofilms, and its sorptive capacity should be able to enhance biological perchlorate removal indirectly by lowering the concentration of oxygen, the competing electron acceptor, through chemisorption. Application of sorptive support media is advantageous for biofilm reactors exposed to transient operating conditions, such as variable influent DO levels, reactor backwashing, and periods without electron donor addition (Choi et al. 2008).

The design and development of biological systems for perchlorate removal has evolved over time. In the early 1990s, the AFRL developed the first reactor, a suspended growth system reactor, which treats wastewater generated when high pressure water is used to remove solid fuels from rockets and missiles, a process often termed ‘hog out’ (Attaway & Smith 1994). The CSTR design was tested at pilot scale at Tyndall Air Force Base in Florida and then installed at the Thiokol rocket production facility in Brigham City, Utah, in 1997. The original Thiokol system consisted of two anaerobic CSTRs (6,000 and 2,700 L) and associated equipment for electron-donor feed and pH adjustment, process control, and effluent clarification and discharge. In 2002, two 3,800-L reactors were added to increase capacity and permit the simultaneous treatment of three different effluent streams containing ammonium perchlorate, potassium perchlorate, and mixed nitrates, respectively. A cheese whey and yeast extract mixture was initially used as an electron donor, but it was later replaced with molasses to reduce cost and improve efficiency. The expanded system is capable of treating ∼3,600 kg perchlorate/month from influent concentrations ranging from 4,000 to 5,000 mg/L (achieved by diluting the concentrated wastewater) to effluent concentrations below the minimum reporting level (MRL) for this matrix (∼400 μg/L) (Air Force Research Laboratory 1998). In 2003, Hodgdon Powder Co. constructed a dual anaerobic CSTR system, the second suspended-growth reactor system, at its gunpowder manufacturing facility in Herington, Kansas. This system, consisting of two 9,500-L molasses-fed reactors designed to process 9,500–19,000 L of wastewater daily, is successfully treating gunpowder processing wastes containing perchlorate at influent levels >3,000 mg/L to effluent levels that are below the MRL of ∼20 μg/L.

In 2004, the California Department of Health Services issued a conditional approval of biological removal of perchlorate from drinking water sources using fixed bed BAC reactors.

PBRs for perchlorate treatment have also been tested in the laboratory (Logan & LaPoint 2002; Min et al. 2004) and a few pilot-scale studies have also been completed with the most extensive pilot testing being at the Texas Street Well Facility in Redlands, California (Min et al. 2004). The PBR design has not yet been applied at full scale for perchlorate treatment, although tests with flow rates as high as 76 L/min have been conducted (Xu 2003). All laboratory and pilot tests show that the PBR can effectively remove perchlorate to a non-detectable level. However, for it to be practical at full scale, a reliable method is needed to control the biomass inventory within the reactor over long periods, while maintaining perchlorate levels below applicable standards.

In addition to biomass inventory, biological treatment has been faced by the challenge of contact between micro-organisms in the treatment unit and the final effluent resulting in secondary contamination of treated water. Research shows that this challenge can be overcome by use of an ion exchange membrane bioreactor (IEMB). Using glycerol as a carbon and energy source, a perchlorate concentration as high as 250 mg L−1 was efficiently reduced to chloride. Despite high perchlorate concentrations in the feed rendering the anion exchange membrane significantly less permeable to perchlorate, the presence of bacteria in the bio-compartment significantly increased the flux through the membrane. Results also suggested minimal secondary contamination (<3 mg C L−1) of the treated water with the optimum feed of glycerol (Fox et al. 2014).

PERCHLORATE REDUCING MICRO-ORGANISMS

A number of micro-organisms have been identified that have the capability to reduce both perchlorate and chlorate. Most identified bacterial strains that reduce perchlorate are facultative anaerobes, mostly Gram-negative. Dissimilatory (per)chlorate reduction has also been reported in various strains including denitrifying bacteria (DB). Many nitrate reducing bacteria in pure cultures reduce chlorate and perchlorate (which is usually referred to as (per)chlorate) by means of membrane-bound respiratory nitrate reductases and assimilatory nitrate reductases (Coates & Achenbach 2004). However, not all DBs can reduce perchlorate. Enzymatic reduction of chlorate to chlorite by nitrate reductase occurs as a competitive reaction between nitrate and chlorate in certain DBs.

Several other microbial isolates have also been obtained that are capable of biodegrading perchlorate through cell respiration, but few of these have been individually tested for perchlorate removal to the required low levels of less than 18 μg/L.

Kim & Logan (2001), while studying microbial reduction in pure and mixed culture PBRs, found that perchlorate can be reduced approximately from 20 mg/L to non-detectable (<4 μg/L) levels in acetate-fed columns inoculated with Dechlorosoma sp. strain KJ or mixed cultures. Flow into the reactor inoculated with the pure culture was at an initial loading rate of 0.24 cm/min (0.06 gpm/ft2), corresponding to an empty bed contact time (EBCT) of 117 min. This was gradually increased to a maximum of 13.6 cm/min (3.35 gpm/ft2; EBCT = 2.1 min) over 83 days of operation in order to determine a minimum EBCT for complete perchlorate removal. Hydraulic loading rates for the reactor inoculated with mixed culture were varied from 0.24 to 0.45 cm/min (0.06 to 0.11 gpm/ft2), corresponding to EBCTs of 43–118 min (detention times of 18–51 min).

It was demonstrated that detention times of PBRs can be substantially reduced using the isolate KJ as compared to a mixed culture, but larger concentrations of acetate, an electron donor, are required to reduce perchlorate to the low levels necessary for drinking water. Perchlorate removal to non-detectable levels, according to the study, required a minimum EBCT of only 2.1 min for the column inoculated with KJ, vs. 31 min for the mixed culture column. Acetate was used at a molar ratio of of 2.9 (n = 6) for the mixed culture, while more than twice as much acetate was consumed on average (6.6 ± 2.0, n = 156) by the pure culture (Kim & Logan 2001).

Despite research focus on PRB, bacteria that degrade perchlorate can be broadly divided into four groups. The choice of any one strain or mixed culture for use in a wastewater treatment facility depends on the characteristics of the wastewater (nitrate, ammonium and salt content, dissolved oxygen level), available carbon sources/electron donors and climatic conditions of the facility location, or energy considerations to maintain the required operation parameters for a given bacteria. These groups are given below.

PRB

These can reduce both perchlorate and chlorate. The majority fall into two distinct monophyletic subgroups: Dechloromonas and Dechlorosoma and most were identified by Achenbach et al. (2001) as a β-subclass of Proteobacteria. In the analysis it was shown that the majority of the PRB in the Rhodocyclus assemblage that form the two distinct monophyletic subgroups, namely, Dechloromonas and Dechlorosoma, fall under the β-Proteobacteria subclass of the Proteobacteria in the 16S rDNA sequence whereas the W. succinogenes strain HAP-1 exclusively represents the ε-Proteobacteria subclass.

They include but have not been limited to: Dechlorosoma, which has recently been renamed Azospira based on the very high (99.9%) 16S rRNA gene sequence identity between the type strain Dechlorosoma suillum and Azospira oryzae; and Dechloromonas such as the hydrogen utilizing strains Dechloromonas sp. JM isolated from hydrogen utilizing autotrophic consortium but incapable of utilizing carbohydrates as an electron donor (Miller & Logan 2000), and Dechloromonas sp. JDS5 and Dechloromonas sp. JDS6 isolated from a perchlorate-contaminated site (Shrout et al. 2005).

Most PRB are facultative anaerobes except for Wolinella succinogenes HAP 1 (Wallace et al. 1996) and Dechlorospirillum anomalous strain WD, which are micro-aerophilic (Coates et al. 1999). All PRB are strict respires and require an e− acceptor for growth. Most are heterotrophic micro-organisms that need a carbon source and can utilize alternate e− acceptors such as O2, nitrate, and chlorate in preference to perchlorate. Most PRB prefer neutral pH and mesophilic temperature. The majority have been isolated under facultative anaerobic conditions using streak plate or shake tube method with perchlorate or chlorate as the e− acceptor. All perchlorate reducers completely reduce perchlorate to O2 and Cl without accumulation of chlorate, chlorite, and O2. Figure 6 shows a microscopic view of Wolinella succinogenes HAP 1.

Figure 6

Microscopic view of Wolinella succinogenes HAP 1.

Figure 6

Microscopic view of Wolinella succinogenes HAP 1.

Chlorate reducing bacteria

These can reduce only chlorate but not perchlorate. Examples include Ideonella dechloratans (Malmqvist et al. 1994), Pseudomonas chloritidismutans strain ASK-1 (Wolterink et al. 2003, 2005), and Alicycliphilus denitrificans strain BC (Weelink et al. 2008).

High chlorate accumulating perchlorate reducing bacteria

These can reduce both perchlorate and chlorate with transient accumulation of chlorate which can be utilized by conventional PRB and chlorate reducing bacteria (CRB) in a syntrophic association. Examples include Dechloromonas PC1 (Nerenberg et al. 2006) and Dechlorosoma sp. HCAP-C (Dudley et al. 2008).

DB

These are not significant players in perchlorate reduction in nature, since perchlorate reduction is not a preferred energy-yielding pathway of denitrifiers. Although DB are able to reduce chlorate, the reaction is not coupled with growth. These bacteria are not likely to grow on chlorate because of the accumulation of toxic chlorite after reduction of the chlorate by the nitrate reductase, which prevents growth. Examples include Rhodobacter capsulatus, Rhodobacter sphaeroides (Roldan et al. 1994), halophilic archaea Haloferax denitrificans, Paracoccus halodenitrificans and A. denitrificans strain BC.

Requirements of PRB

Electron acceptors

PRB have been reported to utilize inorganic e− acceptors such as nitrate, bromate, chlorate, and O2 in preference to perchlorate. Fortunately, if chlorate reducing bacteria are present in the culture, chlorate will not pose a great hindrance since the CRB will reduce chlorate though not perchlorate. In perchlorate-contaminated drinking water, the dominant competing electron acceptors are typically oxygen and nitrate. Consequently, significant perchlorate reduction can only occur after complete removal of nitrate and O2. When perchlorate is the sole e− acceptor, increasingly higher reduction is observed.

Many studies on simultaneous reduction of perchlorate and nitrate have been carried out and all point to the above conclusion (Roldan et al. 1994; Okeke et al. 2002; Cang et al. 2004; Lehman et al. 2008; Ricardo et al. 2012). A study on simultaneous perchlorate and nitrate reduction by a mixed microbial culture in suspension showed that the nitrate reduction rate was 35 times higher than the maximum perchlorate reduction rate. While investigating the biological degradation of nitrate and perchlorate using a mixed anoxic microbial culture and ethanol as the carbon source, it was found that perchlorate reduction was inhibited by nitrate, since after nitrate depletion the perchlorate reduction rate increased by 77% (Ricardo et al. 2012). It was also shown that under ammonia limiting conditions, the perchlorate reduction rate decreased by 10%, whereas the nitrate reduction rate was unaffected. Ammonium ions higher than 0.4% have been found to significantly affect perchlorate reduction of ion exchange regenerant brines.

High concentrations of salts such as sodium bicarbonate or sodium phosphate have an inhibitory effect on the growth of the perchlorate reducers, hence inhibiting biodegradation of perchlorate. Cang developed two cultures capable of degrading perchlorate and nitrate in high salt solutions from marine inoculums. The growth conditions to maintain these cultures in a healthy state required the maintenance of strictly anaerobic conditions and the addition of trace metals, Na2S and phosphate (Cang et al. 2004).

PRB do not utilize other inorganic e− acceptors such as sulfite, sulfate, selenate, Mn-IV (except strain GR-1), and Fe-III, and most of the PRB are unable to use organic electron acceptors. Therefore, their presence in the wastewater has minimal or no effect on the rate of biodegradation of perchlorate by the PRB as is the case for nitrate, chlorate, and oxygen.

Electron donors

PRB can use a wide variety of organic (ethanol, fatty acids, and vegetable oils) and inorganic e− donors (Ju et al. 2008). Acetate is the most commonly used single organic e− donor. Other organic e− donors include acetic acid, lactate, pyruvate, casamino acids, fumarate, succinate, methanol, ethanol, fructose, cellobiose, mannose, xylose, pectin, n-alkanes, 1-hexene, and liquefied petroleum gas. The majority of PRB are, however, unable to use carbohydrates, benzoate, catechol, glycerol, citrate, and benzene.

Although removal has been reported to be successful in studies using organic donors, the organic residual is a concern because it could stimulate bacterial growth in water distribution systems and interfere with chlorination processes, producing disinfection byproducts.

Inorganic electron donors can overcome the disadvantages of organic substrates, and thus are currently the focus of study for biological reduction of . Inorganic e− donors utilized by some PRB include H2, H2S, soluble and insoluble ferrous (Fe-II) iron, ZVI, elemental sulfur (S) and (Son et al. 2006, 2011; Ju et al. 2008; Ahn et al. 2011).

It has been recently shown (Ahn et al. 2014) that pretreatment of the army's insensitive melt-pour explosive, PAX-21, production wastewater with ZVI can convert the energetic compounds present in PAX-21 to products that can serve as electron donors for PRB. PAX-21 mainly contains ammonium perchlorate, RDX, and 2,4-dinitroanisole (DNAN). ZVI reduction experiments showed that DNAN was completely reduced to 2,4-diaminoanisole and RDX was completely reduced to formaldehyde. Anaerobic batch biodegradation of the ZVI-treated PAX-21 wastewater resulted in removal of 30 mg L−1 perchlorate to non-detectable levels within 5 days. Formaldehyde was the primary electron donor for perchlorate respiring bacteria, affirming that integrated iron-anaerobic bioreactor systems can be effective and cost-effective in the biological treatment of perchlorate in army PAX-21 production wastewater (Ahn et al. 2014).

More reliable and effective perchlorate removal can also be achieved by use of the calcium ion. Large amounts of Ca2+ have been reported to result in earlier initiation and faster completion of perchlorate removal. Ca2+ can delay pH increase by combining with OH and consequently extends biodegradation time. If ZVI and Ca2+ co-exist in a system, empty bed residence time can be decreased, and more reliable and effective perchlorate removal performance achieved (Liang et al. 2014).

Nutritional requirements

Microbial organisms responsible for perchlorate reduction need a carbon source just like most micro-organisms. Various substances have been used as carbon sources in biological removal of perchlorate, and the choice of any one carbon source depends on the wastewater characteristics, availability, and the cost implications.

Several PRB require trace metals, molybdenum, iron, and selenium. Lack of vitamins or trace minerals in the growth/isolation medium causes a visible decrease in the degradation rate. Kucharzyk et al. (2012), in a study on maximizing microbial degradation of perchlorate using a genetic algorithm, noted that in the case of Dechlorosoma sp. strain KJ, when 1 mL/L of trace minerals and 0 mL/L of vitamins were applied the degradation rate was only 7.5 mg/L/min. This may be caused by the lack of microelements, such as molybdenum, which are important for perchlorate degradation (Kucharzyk et al. 2012). Protein nutrients such as brewer's yeast, cottonseed protein, or cheese whey and molasses have also been used as nutrient sources (Okeke & Frankenberger 2005). Yeast extract has a stimulatory effect on the growth of these bacteria.

These micro-organisms are capable of utilizing various components of the organic matter in wastes generated from agro-industrial processes as a carbon source for growth and for synthesis of cellular biomass as well (Okeke & Frankenberger 2005). This would go a long way in reducing the costs of substrate requirements during large-scale wastewater treatment.

FACTORS AFFECTING BIODEGRADATION OF AMMONIUM PERCHLORATE

Several studies on biodegradation of perchlorate have been carried out and reports on isolation of PRB published. However, conditions for optimal performance of the PRB have not yet been well understood. The performance of PRB under electron donor limited conditions, oxygenated conditions, and chemically oxidizing conditions needs extensive research in order to optimize microbial perchlorate reduction. Some of the already studied factors affecting performance of PRB are shown in Figure 7.

Figure 7

Factors influencing microbial perchlorate reduction (Bardiya & Bae 2011).

Figure 7

Factors influencing microbial perchlorate reduction (Bardiya & Bae 2011).

Oxygen

Oxygen hinders perchlorate and chlorate reduction by both pure and mixed cultures because PRB utilize oxygen as an electron donor in preference to perchlorate, as mentioned earlier. This is also applicable to dissolved oxygen levels. For instance a 12-h exposure of 6–7 mg L−1 dissolved O2 to a suspended culture of Azospira sp. KJ caused inhibition of perchlorate reduction even after complete removal of the O2. Severe inhibition has also been reported with micro-aerophilic W. succinogenes HAP 1, D. anomalous strain WD and strains JDS5 and JDS6 (Wallace et al. 1996; Coates et al. 1999). Simultaneous reduction of perchlorate and oxygen is therefore necessary to ensure anaerobic conditions essential for the proper activity of the perchlorate reductase. A sufficient supply of an electron donor is therefore imperative for continued perchlorate reduction in oxygenated conditions. This is an added cost to the operation and maintenance of a wastewater treatment system, hence there is a need for cheaper and available electron donors. Biological perchlorate reduction can also be enhanced by use of chemisorption using granulated active carbon as mentioned earlier in the section Biodegredation of ammonium perchlorate (Choi et al. 2008). Using a sorptive biofilm support medium in the reactors can enhance biological perchlorate removal under dynamic loading conditions. As shown in the reactions below, in a redox reaction to account for the reduction of oxygen produced by dismutation, 8 mole of electrons are required to reduce 1 mole of perchlorate to chloride and the produced oxygen to water (Sawyer et al. 2003). 
formula
Reduction and dismutation of perchlorate 
formula
Reduction of oxygen 
formula
Complete reduction of perchlorate 
formula

aCalculated for pH = 7 (Sawyer et al. 2003).

This remains so even in the case of oxygen production due to microbial cell activity of perchlorate grown bacteria. When perchlorate grown bacteria are used in perchlorate reduction, the addition of chlorite yields oxygen outside the cell. However, the organisms appear to be capable of ridding themselves of oxygen produced (even in pure culture) to protect oxygen sensitive perchlorate reducing enzymes, presumably using c-type cytochromes (Rikken et al. 1996).

Perchlorate reduction by mixed lactate (electron donor) enrichment culture (LEC) cannot be slowed by the addition of oxygen in the presence of a sufficient electron donor quantity. Experiments conducted where oxygen was added to active perchlorate degrading microcosms that were designed such that electron donor would not be limiting, as lactate (400 mg/L) was supplied to the microcosms in sufficient quantity to completely reduce the added perchlorate plus 15.8 mg oxygen, showed that, after 8 h incubation with perchlorate followed by room air being injected into the microcosm headspace, the addition of 0.28 mg, 1.1 mg, 2.8 mg, and 5.6 mg oxygen (1 mL, 4 mL, 10 mL, and 20 mL room air, respectively) did not adversely affect perchlorate degradation (Shrout & Parkin 2006). This indicates that perchlorate reduction in a diverse, in situ, bacterial environment can still take place in the presence of molecular oxygen with a sufficient supply of an electron donor. However, if environmental conditions are more oxidized (indicated by a higher redox potential), the rate and extent of perchlorate degradation will be decreased.

Nitrate

The effect of nitrate on perchlorate reduction appears to be more complex than that of oxygen. Close similarity in the reduction potential of the pair (E◦ = 1.25 V, with the pair (E◦ = 1.28 V) makes nitrate an excellent competitor to perchlorate. Consequently, several PRB differ significantly in their response toward the two e− acceptors. The simultaneous reduction of perchlorate and nitrate and the sequential reduction of the two e− acceptors have been extensively reported in the literature both for the pure and enriched cultures. The literature indicates that the two can be simultaneously reduced in the presence of electron donors, but the perchlorate reduction rate decreases slightly in the presence of nitrate. Among the pure cultures, D. agitata strain CKB, W. succinogenes HAP-1 and Perc1ace can reduce nitrate and perchlorate simultaneously, but only Perc1ace can grow with nitrate reduction (Okeke & Frankenberger 2003; Choi & Silverstein 2008). In the majority of cases, the presence of nitrate causes a longer lag in perchlorate reduction, and perchlorate reduction starts only after complete removal of nitrate.

Recent research indicates that absolute perchlorate reduction in a wastewater treatment facility faced with a high concentration of nitrate and sulfate can be achieved using a two-staged hydrogen-based membrane biofilm reactor (MBfR) system (Figure 8) (Zhao et al. 2013) and also by using enzyme-based technologies (Figure 9) (Hutchison et al. 2013).

Figure 8

Two-staged MBfR for perchlorate reduction in the presence of nitrate and sulfate (Zhao et al. 2013).

Figure 8

Two-staged MBfR for perchlorate reduction in the presence of nitrate and sulfate (Zhao et al. 2013).

Figure 9

Perchlorate reduction using free and encapsulated Azospira oryzae enzymes in the presence of nitrate (Hutchison et al. 2013).

Figure 9

Perchlorate reduction using free and encapsulated Azospira oryzae enzymes in the presence of nitrate (Hutchison et al. 2013).

In Figure 8, the surface loading was controlled in each stage and with an equivalent surface loading larger than 0.65 ± 0.04 g N/m2-day, the lead MBfR removed about 87 ± 4% of NO3 and 30 ± 8% of . This reduced the equivalent surface loading of to 0.34 ± 0.04–0.53 ± 0.03 g N/m2-day for the lag MBfR, in which was reduced to non-detectable levels. reduction was eliminated without compromising full reduction using a higher flow rate that gave an equivalent surface loading of 0.94 ± 0.05 g N/m2-day in the lead MBfR and 0.53 ± 0.03 g N/m2-day in the lag MBfR. The lead MBfR biofilm was dominated by DB, Dechloromonas, Rubrivivax, and Enterobacter, whose main function was to reduce nitrate, whereas the lag MBfR was dominated by PRB, Sphaerotilus, Rhodocyclaceae, and Rhodobacter, with the main function of perchlorate reduction as nitrate loading being small (Zhao et al. 2013).

Figure 9 shows the removal of perchlorate in the co-occurrence of nitrate using cell-free bacterial enzymes as biocatalysts. In this study, crude cell lysates and soluble protein fractions of Azospira oryzae PS, as well as soluble protein fractions encapsulated in lipid and polymer vesicles were used. Perchlorate was removed by the soluble protein fraction at higher rates than nitrate and perchlorate reduction even in the presence of 500-fold excess nitrate (Hutchison et al. 2013).

Salinity

Perchlorate reduction is drastically affected at salt concentrations as low as 1%, and is completely inhibited at salt concentrations above 4% (Coates et al. 1999). Microbiological treatment of perchlorate containing solutions has not been reported thus far to occur at the high salinities typical of perchlorate-contaminated ion exchange brines. A perchlorate degrading isolate obtained by Logan et al. (2001) had an optimal salinity of 1% and grew only at <2% NaCl. Non-salt-tolerant PRB is completely inhibited at salinities over 2–4% (Michaelidou et al. 2000; Gingras & Batista 2002), but salt-tolerant PRB that is acclimated to high salinity has been shown to reduce perchlorate at appreciable rates in high salinity, although the rates are indirectly proportional to the salt concentrations.

Owing to the high ionic strength of industrial wastewaters containing high NaCl concentrations, it is very difficult to remove perchlorate as high ionic strength of wastewater has been reported to hinder the perchlorate reduction activity of perchlorate reducers. The inhibitory effects of high ionic strength can be dealt with by (1) diluting the wastewater until the inhibitory effect dissipates and (2) using PRB strains that are tolerant of high ionic strength. However, the dilution process is operation and capital intensive, leaving the use of salt tolerant PRB the more economical option. In a recent study on perchlorate reduction using salt tolerant bacterial consortia, it was found that although the perchlorate reduction rates decreased with increasing NaCl concentration, salt tolerant-PRBl consortia could reduce perchlorate to 75 g-NaCl L−1 (Ryu et al. 2011).

Temperature

Perchlorate reduction has been reported to occur over a wide range of temperatures (10–40 °C), however, optimal reduction proceeds between 28 and 37 °C. The temperature range for bacteria growth of W. succinogenes HAP 1 was 20–45 °C, with an optimum at 40 °C. Perchlorate reduction by the perchlorate respiring bacterium Perc1ace was achieved in the temperature range of 20–40 °C, and with optimum activity at 25–35 °C (Okeke & Frankenberger 2003).

pH

Most PRB require a neutral pH of around 6.8–7.2 for growth and optimal perchlorate reduction, although perchlorate reduction occurs throughout the pH range from 5.0 to 9.0. Rates of perchlorate removal by a unit mass of bacteria have been reported to be significantly different at various pHs, with a maximum rate at pH 7.0 (Wang et al. 2008). The optimal PRB growth for W. succinogenes HAP 1 occurs at a pH of 7.1, whereas the optimal perchlorate reduction rate in a mixed culture occurs at a pH of 7.0. Table 5 shows a summary of pH ranges for both pure and mixed heterotrophic PRB (Wang et al. 2008).

Table 5

pH ranges for PRB

  PH range for PRB
 
  
Pure or mixed culture Perchlorate reduction Bacteria growth Reference 
HAP-1 N/A 6.5–8.0, optimum 7.1 Wallace et al. (1996)  
Perclace 6.5–8.5, optimum 7.0–7.2 N/A Okeke & Frankenberger Jr (2003); Coates & Achenbach (2004)  
Mixed culture 5.0–9.0, optimum 7.0 N/A Wang et al. (2008)  
CKB N/A 6.5–8.5, optimum 7.5 Herman& Frankenberger (1999)  
Acinetobacter N/A 6.0–7.5, optimum 6.8–7.2 Stepanyuk et al. (1992)  
Mixed culture 6.6–7.5, optimum 7.1 Wide range Attaway & Smith (1993)  
  PH range for PRB
 
  
Pure or mixed culture Perchlorate reduction Bacteria growth Reference 
HAP-1 N/A 6.5–8.0, optimum 7.1 Wallace et al. (1996)  
Perclace 6.5–8.5, optimum 7.0–7.2 N/A Okeke & Frankenberger Jr (2003); Coates & Achenbach (2004)  
Mixed culture 5.0–9.0, optimum 7.0 N/A Wang et al. (2008)  
CKB N/A 6.5–8.5, optimum 7.5 Herman& Frankenberger (1999)  
Acinetobacter N/A 6.0–7.5, optimum 6.8–7.2 Stepanyuk et al. (1992)  
Mixed culture 6.6–7.5, optimum 7.1 Wide range Attaway & Smith (1993)  

CONCLUSION

Perchlorate, being poorly reactive and highly soluble in water, can persist in the environment for long periods of time. Conventional physical and chemical water and wastewater treatment processes are inapplicable for the removal of the perchlorate ion. Treatment options that have been used to remove the perchlorate ion have necessitated additional steps to treat or dispose of the concentrated perchlorate residual waste stream that is generated, leaving biological treatment as the only viable method that can completely remove from wastewater effluents.

To curb challenges limiting wide acceptability of biodegradation, effective biological treatment of perchlorate wastewater in large-scale applications such as army munition production can be achieved through integration of two or more treatment options, such as inclusion of tailored GAC in biological reactors, whose main disadvantage is the long-term need for further treatment of explosive-laden spent carbon or disposal by landfill/incineration.

Biomass inventory and secondary contamination of treated water may be addressed through the integration of ion exchange with a membrane bioreactor as an IEMB system. Electron donor availability and electron competition can be reduced by pretreatment of army PAX-21 production wastewater with ZVI before biodegradation, as ZVI pretreatment breaks energetic compounds into compounds that can be used by PRB as electron donors. ZVI can be combined with Ca2+ for a more reliable, fast, and effective perchlorate removal efficiency, as the Ca2+ reacts with OH delaying the pH increase, hence extending biodegradation time. Electron competitors such as sulfate and nitrate can also be dealt with by using two-staged hydrogen-based MBfR systems and also by using enzyme-based technologies.

However, further research and pilot-scale application on the aforementioned recommendations is a necessity.

ACKNOWLEDGEMENTS

The present research was supported by the Major Science and Technology Program for Water Pollution Control and Treatment (2012ZX07201002-6), National Scientific Funding of China (51378003), International Cooperation Project (2013DFG92600), Beijing Higher Education Young Elite Teacher Project.

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