A critical review of ionizing radiation technologies for the remediation of waters containing Microcystin-LR and M. aeruginosa

Abstract

Harmful algal and cyanobacterial blooms pose threats to human and ecological health due to their release of hazardous toxins. Microcystin-LR (MC-LR), a potent hepatotoxin, is the most prevalent cyanotoxin found in freshwater blooms. Although produced by many species of cyanobacteria, Microcystis aeruginosa is most commonly associated with MC-LR production. These blooms are increasing in occurrence in lakes, ponds, and other surface waters and, therefore, require efficient treatment methods to be removed from water supplies. Ionizing radiation technologies offer promising approaches for the removal of organic pollutants in water, including cyanotoxins and cyanobacteria. Gamma irradiation for the degradation of cyano-bacteria and toxins is effective for overall MC-LR degradation as well as reducing cell concentrations. However, gamma irradiation technology involves use of radioactive isotopes and, therefore, may not feasible commercially from a security perspective. Electron beam (eBeam) irradiation technology, which relies on regular electricity to generate highly energetic electrons, is able to achieve the same results without the confounding challenges of radioactive isotopes and related security issues. In this critical review, the current state of the science concerning the remediation of MC-LR and M. aeruginosa with ionizing radiation technologies is presented and future necessary research is discussed.

1. Introduction

The increasing occurrence of harmful algal blooms (HABs) is contributing to the decrease of an already dwindling water supply world-wide. HABs occur due to ecosystem imbalances where large colonies of algae grow out of control and release odor and taste compounds as well as harmful toxins into the surrounding waterbody (Sharma et al., 2012). Increased occurrence of HABs can be attributed to rising temperatures and escalating anthropogenic nutrient pollution (namely nitrogen [N] and phosphorus [P]) in affected waterbodies (Paerl and Otten, 2013). However, no one factor has been identified as the cause of these blooms (Preece et al., 2017). The resulting eutrophication prompts accelerated algal growth developing into thick mats of green biomass. Human exposure primarily occurs through ingestion of contaminated water during recreational activities or via insufficiently treated drinking water.

The main culprits of freshwater blooms are not eukaryotic algae, but cyanobacteria (previously classified as blue green algae). Cyanobacteria differ from heterotrophic bacteria in that they are the only phylum of bacteria that use oxygenic photosynthesis as their main form of energy. Certain cyanobacteria are also able to reduce nitrogen and carbon in aerobic conditions which has aided in their evolution and their ability to exist in marine, freshwater, and dry habitats (Gault and Marler, 2009). Many species also produce various toxic secondary metabolites (cyanotoxins). It is estimated that approximately 75% of cyanobacterial blooms exhibit toxicity (Preece et al., 2017). These toxins are generally classified by mechanism of action or chemical structure (Ferrão-Filho and Kozlowsky-Suzuki, 2011). Primarily responsible for neuro- or hepato-toxicity, Table 1 summarizes common toxins, their producers, and their mechanisms of action (Wiegand and Pflugmacher, 2005; Ferrão-Filho and Kozlowsky-Suzuki, 2011).

Table 1.

Common cyanotoxins, main producing genera, and primary mechanism of action.

Toxin	Toxin Producing Genera	Associated Toxicity	Primary Mechanism of Action
Microcystins	Anabaena, Microcystis, Plankthotrix	Hepato-	Inhibition of protein phosphatases (PP1 and PP2A)
Nodularins	Nodularia	Hepato-	Inhibition of protein phosphatases (PP1 and PP2A)
Saxitoxins	Aphanizomenon, Anabaena, Cylindrospermopsis, Lyngbya	Neuro-	Binding and blocking the sodium channels in neural cells
Anatoxins	Anabaena, Aphanizomenon, Cylindrospermopsis, Microcystis, Oscillatoria, Planktothrix	Neuro-	Binding irreversibly to the nicotinic acetylcholine receptors
Anatoxin – a	Anabaena	Neuro-	Inhibition of Ach-esterase activity
Cylindrospermopsin	Anabaena, Aphanizomenon, Cylindrospermopsis, Raphidiopsis, Umezakia	Hepato-	Inhibitor of protein biosynthesis cytogenetic damage on DNA
Lipopolysaccharide	Various	Various	Potential irritant; affects any exposed tissue

Adapted from Wiegand and Pflugmacher (2005) and Ferrão-Filho and Kozlowsky-Suzuki (2011).

1.1. Microcystins

Of particular importance to human health are microcystins. Microcystins (MCs) are a group of hepatotoxic cyanotoxins produced by a variety of cyanobacteria including: Microcystis spp., Anabaena spp., and Plankthotrix spp., and to a lesser extent Dolichospermum spp., Geitlerinema spp., Leptolyngbya spp., Pseudanabaena spp., Synechococcus spp., Spirulina spp., Phormidium spp., Nostoc spp., Oscillatoria spp., and Radiocystis spp. (Buratti et al., 2017). Microcystis aeruginosa is the most common cyanobacterial species found in freshwaters worldwide and has been associated with a number of human, livestock, and wildlife poisonings (Tanabe et al., 2018). M. aeruginosa commonly produces Microcystin-LR (MC-LR) which is the most toxic and most prevalent of the over 100 identified variants of MCs (Buratti et al., 2017). All MCs share a common structure including a cyclic heptapeptide containing 3 D-amino acids (alanine, glutamic acid, and mathylaspartic acid), two ‘unusual’ amino acids (N-methyldehydroalanine and 3-amino-9-methoxy-2,6,8-trimethyl-10-phenyldeca-4,6-dienoic acid (ADDA)), and two variable L-amino acids (X and Z) (Sharma et al., 2012). MC-LR contains leucine and arginine in the X and Z positions, respectively, and accounts for 46–99.8% of total HAB microcystin concentrations (Fig. 1) (Vasconcelos et al., 1996). Other less common variants include MC-LA, MC-YR, MC-RR, MC-LF, and MC-LW, however, these variants have varying levels of toxicity less than that of MC-LR.

Fig. 1.

Chemical structure of microcystin-LR.

MC-LR’s biologic activity is attributed to the ADDA moiety and simple stereochemical changes to this group have been shown to drastically reduce the molecule’s toxicity (Tsuji et al., 1994; Song et al., 2006). Mechanistically, MC-LR targets hepatocytes in the liver and enters via active transport through transporters in the organic anion transporting polypeptide (OATP) superfamily (Fig. 2) (Fischer et al., 2005; Valério et al., 2010). After entering the cell, MC-LR binds strongly and irreversibly to serine/threonine protein phosphatases 1 and 2A (PP1 and PP2A) resulting in inhibition of these enzymes (Craig et al., 1996). Given their importance in cell function and cell cycle regulation, inhibition of PP1 and PP2A lead to hyper-phosphorylation of proteins and cytoskeletal filaments which can induce apoptosis. MC-LR ingestion may also result in DNA damage, cell proliferation, and possible tumor promotion (Valério et al., 2010). Acute toxicity can result in liver inflammation, hemorrhaging, and extensive hepatic bleeding. Death may occur due to liver failure at high or prolonged doses.

Fig. 2.

Microcystin-LR mechanism of action. Adapted from Valério et al. (2010). Was published open access in Toxins and is covered under a Creative Commons Attribution License (https://creativecommons.org/licenses/by-nc-sa/3.0/.

Toxic algal blooms have been noted in literature as far back as 1878 when George Francis published “Poisonous Australian Lake” (Francis, 1878). Francis described the presence of thick algae as looking like green oil paint inches deep on Lake Murray, as well as the death of livestock following ingestion of the green muck. Animal lethalities due to MC-LR have since been reported in a variety of species, including: flamingos, cattle, dogs, ducks, laboratory rats and mice, turtles, etc (Mahmood et al., 1988; Fitzgerald and Poppenga, 1993; Gupta, 1998; Matsunaga et al., 1999; Krienitz et al., 2003; Nasri et al., 2008; van der Merwe et al., 2012; Backer et al., 2013). The largest human intoxication event occurred in 1996 when a hemodialysis center in Caruaru, Brazil experienced the largest modern day human microcystin poisoning due to contaminated water used in treatment. 100 patients developed acute liver failure after being exposed to 19.5 μg/L of MC intravenously during treatment. Of these patients, 76 died of the resulting liver failure (Jochimsen et al., 1998; Carmichael et al., 2001). A study that followed soon after correlated chronic sublethal ingestion of microcystins in drinking water with an increased incidence of liver cancer in China (Yu et al., 2001). Recently in 2011, the Kansas Department of Health and Environment received 25 reports of human illness due to cyanotoxins. Seven of the reports were confirmed by follow up and symptoms were consistent with MC exposure. Fortunately, there were no fatalities (Trevino-Garrison et al., 2015).

1.2. Remediation strategies

The increasing health concerns of MCs have prompted research into more effective removal or remediation strategies in water treatment. Current treatment relies on conventional oxidants, mainly ozone, to remove MCs in water. However, these conventional oxidants have varying levels of efficiency, are reliant on specific operational parameters for breakdown of pollutants, and are chemical additives which may create secondary harmful disinfection byproducts (Rositano et al., 2001; Teixeira and Rosa, 2007; Sharma et al., 2012; Fan et al., 2014). Other advanced oxidation processes (AOPs) or advanced oxidation reduction processes (AORPs) could present more efficient and chemical-free approaches for degrading MCs and other pollutants. In particular, ionizing radiation technologies have been proven effective at removing a variety of chemicals and biologics in water systems (Nickelsen et al., 1994; Yu et al., 2009; Praveen et al., 2013; He et al., 2014; Wang et al., 2016).

During ionizing radiation treatment, the radiolysis of water creates both reactive oxidative and reductive species (eq. (1)). The values in brackets represent ‘G-values’ or quantities of each species produced per 100 eV of absorbed energy (Miller, 2005).

$${\text{H}}_2\text{O}⟿[2.6]{\text{e}}_{\text{aq}}^{-}+[0.55]{\text{H}}^{•}+[2.7]{\text{H}}_3{\text{O}}^{+}+[0.7]{\text{H}}_2{\text{O}}_2+[2.6]{\text{HO}}^{•}+[0.55]{\text{H}}_2$$ (1)

Non-specific interaction of the electrons and/or the other reactive species with chemical and biological pollutants result in the degradation of the chemicals and the inactivation of the organisms. In living organisms, these interactions induce single stranded and double stranded DNA strand breaks as well as chemical alterations within the cell.

Given the growing applications of ionizing radiation in environmental remediation (Cooper, Curry and O’Shea, 1998) there is a need to explore the utility of ionizing radiation technology for emerging contaminants such as cyanotoxins and toxin-producing cyanobacterial cells. The primary objective of this review is to describe the current state of the science surrounding the use of ionizing irradiation technologies for the degradation of MC-LR and the inactivation of the cyanobacteria, M. aeruginosa.

2. Methods

Qualitative systematic review techniques were employed in this overview. Existing studies were obtained using a defined set of search terms (see Table S1). The databases utilized included Google Scholar, Science Direct, and PubMed. Articles were excluded if: microcystins were not the target chemical or M. aeruginosa was not the target organism, ionizing radiation was not utilized, or the article was not available in English translation. No studies utilizing x-ray radiation techniques were identified. Qualifying articles were reviewed, and results are summarized. Articles were included regardless of study quality. However, strengths and/or limitations, use of statistics, and bias were discussed for each study.

3. Current research (ionizing irradiation technologies)

3.1. Gamma irradiation

In the 1990s, Wayne W. Carmichael published a series of articles on cyanotoxins and cyanobacterial blooms which emphasized the importance of monitoring cyanotoxins in water (Carmichael et al., 1990; An and Carmichael, 1994; Carmichael, 1994). However, preliminary research started as early as the 1960s on irradiation techniques for the removal of these toxins.

Chronologically speaking, the first paper that investigated the controlling of cyanobacterial populations with ionizing irradiation was Morton and Derse in 1968 (Morton and Derse, 1968). The authors aimed to investigate the use of gamma irradiation to control algal/cyanobacterial culture growth. They investigated five HAB related organisms including: Anabaena circinalis, Aphanizomenon flos aquae, Microcystis aeruginosa, Chlorella pyrenoidosa (green algae), and Gomphonema sp. (diatom). Radiation was performed using a 500 Ci Co⁶⁰ gamma source and samples were dosed based upon exposure times. Concentration of cells in cultures were determined using absorbance at 600 nm on a spectrophotometer. The authors reported that 1–1.5 kGy was necessary to ‘control growth substantially’ for Chlorella, Anabaena, and Microcystis, whereas Gomphonema and Aphamizomenom were more resistant. In particular, M. aeruginosa at bloom level concentrations (approximately 1 × 10⁶ cells/ml) required a dose greater than 1 kGy to achieve an absorbance of 0 after culturing 11 days post-irradiation. The authors noted that there was no observed initial concentration dependence on the dose viability for controlling populations of any species. Further, the authors compared the results of the irradiated cultures to previously published work that treated cultures chemically (algicides or algistats) to prevent algal growth. They concluded that the significant variability seen between species and between cell concentration with various algicides was not seen with gamma treatment. Overall, this initial work by Morton and Derse suggested that cell growth could be prevented in M. aeruginosa and other cyanobacterial species using gamma irradiation.

A study published the following year by Kraus (1969) also used a Co⁶⁰ gamma source for investigation of the resistance of 23 cyanobacterial strains (including M. aeruginosa) to irradiation (Kraus, 1969). In the study, Kraus determined cell viability based on a change in dry weight of pelleted cells due to differing cell morphologies. They also used a radiation-resistant bacterium, Micrococcus radiodurans, and a non-radiation resistant bacterium, Sarcina lutea, in co-culture with cyanobacterial species for comparison of resistance. Cultures were grown in a modified Chu medium with doubled nitrate content to discourage the cyanobacterial production of extracellular polysaccharides. Cell concentrations of irradiated cultures were not provided. Following irradiation, cultures were grown in fresh media for 15–21 days to monitor cell regrowth. Using the LD₉₀ determined from irradiated cultures, the authors categorized cyanobacterial resistance in groups; low resistance (LD₉₀ < 4 kGy), moderate resistance (4 kGy < LD₉₀ < 12 kGy), and high resistance (12 kGy < LD₉₀). M. aeruginosa was considered “sensitive” and placed in the low resistance group. The authors also note that microscopic analysis of cells after ‘moderate’ exposures showed no abnormalities in cell division. After ‘high’ exposures there appeared to be distortion of the cell. However, this data was not provided.

Although recent research has improved methods for studying organism effects, these early studies provided a foundation for irradiation treatment of cyanobacteria and other microorganisms. Presumably unknown at the time that these studies were conducted, was the understanding that the presence of specific chemicals in solution can fundamentally alter the types of reactive species abundant in solution after radiation by virtue of their radical scavenging effects (Khan et al., 2015; Ma et al., 2017). The addition of excess nitrate in media as performed by Kraus may have caused scavenging of aqueous electrons (e_aq⁻) reducing species as well as generation of NO₃²⁻ which may have affected cyanobacterial removal (Wang et al., 2016).

There was a gap of approximately 38 years before the next gamma study was published focusing on M. aeruginosa and microcystins. This was published in 2007 by Zhang et al. (2007). This study focused on the radiolysis of MC-LR and MC-RR by gamma irradiation, as well as the effect of additives on degradation. M. aeruginosa was cultured in BG-11 medium with additives being added directly to the culture medium in those studies. Microcystin concentrations were quantified using an HPLC with a UV diode array detector. The authors reported that degradation of both MCs increased with increasing dose. A dose of 8 kGy resulted in a 98.8% degradation efficiency of MC-LR and a dose of 5 kGy was enough to remove all MC-RR. The authors then calculated D 0.9 values (the required dose to reduce 90% of the initial concentration) resulting from the addition of additives to M. aeruginosa cultures. Presently, these values are referred to as D₁₀ values to denote the reduction of the population by a factor of 10. Na₂CO₃ and H₂O₂ were seen to enhance both MC-LR and MC-RR degradation efficiency, whereas NaNO₂, NaNO₃, and Triton X-100 were seen to inhibit degradation. They suggest that the presence of CO₃²⁻ quenches the H₃O⁺ created by gamma irradiation’s radiolysis of water resulting in an increase in available e_aq⁻. This would suggest a reductive process for MC break-down. Oppositely, the authors also suggest the presence of H₂O₂ could act as a source of hydroxyl radicals in solution promoting more oxidative processes. However, studies have shown that hydrogen peroxide may also act as a scavenger of hydroxyl radical (Henry and Donahue, 2011). Nitrate-containing compounds may react with both e_aq⁻ and H to inhibit the reduction process. Effects of all additives seemed more pronounced at lower irradiation doses.

Song et al. (2009) completed pulse radiolysis experiments and gamma irradiation studies on MC-LR (Song et al., 2009). In these studies, Microcystin was purified from M. aeruginosa cultures and purity was determined through HPLC. Radiolysis was performed using an 8 MeV Titan Beta model TBS-8/16–1S linear accelerator. Gamma irradiation was completed using a Co⁶⁰ source with Fricke dosimetry. Breakdown of MCLR and degradation products were analyzed with LCMS. The authors began by conducting kinetic studies with 2–3 ns pulsed radiolysis at 3–5 Gy. By identifying experimental rate constants, addition of HO^∙ to the unsaturated hydrocarbons on MC-LR including the ADDA moiety benzene (1.03 ( ± 0.03) × 10¹⁰ M⁻¹ s⁻¹), and ADDA moiety diene (10⁹-10¹⁰ M⁻¹ s⁻¹), were determined to be the fastest reactions. Overall, the authors determined the rate constant for the reactions of hydroxyl radicals with MC-LR to be 2.3 ( ± 0.1) × 10¹⁰ M⁻¹ s⁻¹. Hydrogen abstraction was seen to occur 1–2 magnitudes slower than the other reactions (10⁸ M⁻¹ s⁻¹), however, more than 50 reaction sites exist on the MC-LR molecule making it a significant reaction pathway. The authors then modelled hydroxyl radical reactivity in respect to individual amino acids in MC-LR. Overall, a rate constant of 2.1 × 10¹⁰ M⁻¹ s⁻¹ was obtained by summation of these individual reaction sites which was approximately 10% lower than the observed experimental rate constant. They noted this was within experimental error and could be in part due to the exclusion of hydrogen abstraction pathways for the ADDA moiety. Next, Song et al. investigated transformation pathways using gamma irradiation. Samples were saturated with oxygen to encourage the reaction of e_aq⁻ and hydrogen atoms with dissolved oxygen. This in turn produced superoxide anions with much lesser reactivity than hydroxyl radicals. A dose of 1.8 kGy de-graded MC-LR. Further, degradation products corresponding to hydroxylation of the benzene on the ADDA moiety and hydroxyl attack on the diene of the ADDA moiety were identified at 1011 m/z and 1029 m/z, respectively (Fig. 3). Overall, these studies suggest that the degradation products of MC-LR due to eBeam irradiation (as seen in the pulse radiolysis experiments) could undergo similar oxidative reactions to that of gamma irradiation. Further, the breakdown of the molecule appears to be due to oxidative reactants rather than reductive as in the absence of reductive species, MC-LR degradation products were still identified.

Fig. 3.

Reaction products for hydroxyl radical reaction with the A) benzene group and B) diene group of ADDA moiety of MC-LR. Reprinted (adapted) with permission from Song, Weihua, et al. “Radiolysis studies on the destruction of microcystin-LR in aqueous solution by hydroxyl radicals.” Environmental Science & Technology 43.5 (2009): 1487–1492. Copyright (2009) American Chemical Society.

Zheng et al. (2012) investigated the use of gamma irradiation to remove M. aeruginosa in water (Zheng et al., 2012). Various additives were also tested for their influence on cell removal. Cyanobacteria cultures were grown in BG-11 media and irradiated with a Co⁶⁰ gamma source. Cell growth was monitored using chlorophyll-a concentrations determined at 665 nm absorbance and carotenoid content determined at 615 and 652 nm absorbance. The antioxidants superoxide dismutase (SOD) and peroxidase (POD) were also monitored as a marker for oxidative stress in M. aeruginosa. Additives used included CH₃OH, thiourea, and NO ⁻₃ and pH effects were measured. The authors report that a decrease in chlorophyll-a content was seen with increasing dose five days post-irradiation. Change in culture color was visually observed and a dose of 9 kGy resulted in a 98% chlorophyll-a removal efficiency. Similarly, carotenoids were greatly affected by irradiation with only 0.8% of control carotenoid content remaining after 9 kGy of irradiation. pH was also seen to impact removal efficiency with an increasing pH resulting in a decreased M. aeruginosa removal. The authors suggest this result could be due to H^∙ readily reacting with OH⁻ in alkaline conditions to produce more e_aq⁻. Further, recombination of e_aq⁻ and OH^∙ reduces hydroxyl radical concentrations and therefore decreases oxidative reactions in solution. However, M. aeruginosa has not shown to be adapted to high or low pH, suggesting that some observed cell death could have been due to culture pH and not only gamma irradiation. The authors then discussed the effect of additives on M. aeruginosa removal. CH₃OH addition resulted in a slight increase in chlorophyll-a content with increasing dose suggesting that removal of M. aeruginosa is reliant on OH^∙. Addition of thiourea also slightly increased the chlorophyll-a content suggesting H^∙ and e_aq⁻ are also involved in M. aeruginosa removal. Finally, addition of NO₃⁻ also slightly decreased the removal efficiency of M. aeruginosa. Unfortunately, due to a lack of statistical significance, it was unclear which primary species were responsible for cell removal. The authors then reported changes in SOD and POD in irradiated cells. At low doses (2–5 kGy), SOD and POD activity was increased, caused by an increase in oxidative stress. High doses (6–9 kGy) reduced SOD and POD activity. The authors did not offer any explanation of the observed effects. However, at high doses, there may be increased DNA damage resulting in a dampened stress response and therefore reduced activity of SOD and POD. Lastly, Zheng et al. presented SEM images of M. aeruginosa cells irradiated at 9 kGy. Irradiated cells appeared to have depressions across the cell surface suggesting that irradiation could be affecting cellular morphology as well.

3.2. Electron beam irradiation

The first reported study to investigate eBeam as a possible treatment technique was a pilot project completed by Ho Kang in 2004 as a part of the International Atomic Energy Agency’s (IAEA) coordinated research project on Remediation of Polluted Waters and Wastewater by Radiation Processing (Kang, 2004). The study briefly investigated the use of eBeam to damage a variety of algal and cyanobacterial species including: Chlorella sp., Scenedesmus sp., Microcystis sp., Anabaena sp., Oscillatoria sp., Prorocentrum micans, Prorocentrum mininum, Scrippsiella trochoidea, Lingulodinium polyedra, and Cochlodinium polykrikoides. A Russian ELV-4 model electron beam accelerator was used to treat samples and samples were dosed from 1 to 10 kGy. The author reported that for Microcystis, a 40% reduction in photosynthetic activity was seen following a 3 kGy dose. A similar reduction was seen in marine red algae at approximately 1 kGy. However, no methods for chlorophyll-a measurement are mentioned in the study. The author also noted that after irradiation, freshwater algae leached soluble proteins from cells and a ‘biopolymeric substance’ that lead to bioflocculation of cells within two days of eBeam treatment. However, there was no data in the paper to support this claim.

Only two other papers have been identified in literature relating to eBeam treatment of M. aeruginosa and microcystin. The first was published by Liu et al. (2014) and investigated the effect of eBeam irradiation on M. aeruginosa (Liu et al., 2014). M. aeruginosa cultures were grown in BG-11 media at a pH of 7.5. Samples were irradiated in a glass petri dishes in 100 ml volumes with a low energy linear accelerator (1.0 MeV and 1.0 mA). Chlorophyll-a was extracted using 90% acetone to determine photosynthetic ability and cell concentrations were monitored through optical density (OD₆₈₀) measurements at 680 nm. The authors reported a removal efficiency based upon chlorophyll-a content of 43%, 83%, 86%, 91%, and 84% for doses of 1–5 kGy, respectively. OD₆₈₀ measurements decreased 34%, 71%, 74%, 85%, and 80% for doses 1–5 kGy, respectively. However, it was unclear if replicates were completed in the experiments. Additionally, the authors studied the effect of eBeam on cell morphology using a JEM-1230 transmission electron microscope (TEM). Although images were provided for dosed cells, the authors only suggest that eBeam treatment can cause damage to integrity and morphology of M. aeruginosa cells. Next, they reported on decreases in photosynthetic rate of cultures following > 2 kGy eBeam dose. Here, no method details were included to suggest how photosynthetic rates were obtained. The authors provide data on changes in SOD and POD enzymatic activity. Up to 7 days post irradiation, the authors saw an increase in POD activity, followed by a decrease up to 11 days. Similarly, up to 5 days post irradiation, the authors saw an increase in SOD activity, followed by a decrease up to 11 days. These results were attributed to oxidative damage caused by eBeam irradiation. Unfortunately, the experimental methods used for these studies were not disclosed.

The final study was published by the same group (Liu et al., 2015) and investigated the use of eBeam irradiation to control microcystin concentrations (Liu et al., 2015). Similar to the previous publication, chlorophyll-a content and OD₆₈₀ were monitored and irradiation was completed in glass petri dishes using a low energy beam (1.0 MeV and 1.0 mA). After irradiation cultures were analyzed for intracellular and extracellular microcystin content. MC concentration was determined using an ELISA kit. The authors reported that for both intra- and extracellular MC-LR content, there was an increase in MC at approximately 4 days post irradiation for doses > 1 kGy, followed by a sharp decrease up to 12 days. Although there is an increase, the authors conclude that an appropriate dose of eBeam irradiation can inhibit MC production. Liu et al. further presents total MC concentration data and reports that 37.2%, 60.8%, 59.6%, 60.2%, and 72.1% MC decreases were observed for doses of 1–5 kGy, respectively. It is unclear at what amount of time post irradiation these samples were taken. Finally, the authors tried to correlate MC concentration with algal growth using chlorophyll-a content. They found that chlorophyll-a content seemed to increase with increasing MC concentration for both control and irradiated samples. As with the previous publication, there was no mention of experimental replication.

4. Future research directions

The studies published to date illustrate promising results for the use of ionizing radiation technologies for the breakdown and removal of MC-LR and M. aeruginosa in water.

Since gamma irradiation technology predates electron beam irradiation technology, there have been a larger volume of studies investigating its effectiveness for pollutant and organism treatment. Gamma irradiation could be a feasible treatment option for MC-LR and M. aeruginosa in water. However, the concept of utilizing radioactive cobalt-60 isotopes for the remediation of cyanotoxins, as well as other emerging water contaminants, is untenable from both a technology perspective as well as a homeland security perspective. Cobalt-60 is a high security material that requires extensive protection in transportation, handling, storage, and disposal. Agencies around the world are actively trying to reduce the commercial use of this technology (Lubenau and Strom, 2002; National Science and Technology Coucil et al., 2016; Chou et al., 2018). Therefore, the need for radioactive materials makes widespread usage of gamma irradiation technology doubtful in today’s security and environmentally conscious world.

There is a growing body of literature highlighting the value of eBeam technology for environmental remediation applications (Nickelsen et al., 1994; Chaychian et al., 1998; Yu et al., 2009; Praveen et al., 2013; He et al., 2014; Wang et al., 2016). The preliminary data for eBeam irradiation as a water treatment technology for MC-LR and M. aeruginosa is promising. Electron beam technology utilizes a linear accelerator to generate its highly energetic electrons from commercial electricity. Therefore, this is a switch on and off technology without the security issues associated with radioactive sources. The U.S. Department of Energy has recognized the potential of this technology for energy and environmental applications (U.S. Department of Energy, 2015). Compact high energy (10 MeV) and high power (700 kW) linear accelerators capable of treating extremely large volumes of water are commercially available today. However, there is a lack of published articles related to eBeam technology and the degradation of cyanobacteria and their toxins.

There is a critical need for pursuing research on cyanobacteria and cyanotoxin remediation on multiple fronts. One key objective should be to understand the breakdown products associated with MC-LR degradation. It is important to understand the extent of breakdown of the toxin molecule that is achievable at varying eBeam doses and under varying experimental conditions such as pH, total organic carbon content, presence of extraneous biomass, and chemical composition, etc. A deep understanding of the expected toxin breakdown products under varying eBeam doses will help in predicting possible toxicity associated with these by-products. These studies can also shed light on whether the by-products would undergo autolysis or will be metabolized by the microbial community. In vitro and in vivo toxicity studies using the eBeam degradation products are a necessary compliment to the above described studies.

Another research focus should be to understand the cellular effects of eBeam doses on M. aeruginosa and other toxin-producing cyanobacteria. It is important to confirm whether or not the cell undergoes morphological changes. It is now known that in bacterial cells, eBeam irradiation does not cause cell lysis but nevertheless causes inactivation (Jesudhasan et al., 2015). Experimental approaches should include microscopy as well as the use of membrane integrity dyes and nucleic acid stains. Microscopic studies should be performed at specific time points post-eBeam irradiation exposure as well as after incubation at varying conditions. A clear understanding of how the cells respond to varying eBeam doses will help in developing post treatment cell filtration/separation strategies.

A third avenue of research should revolve around understanding how the genome, transcriptome, proteome, and the metabolome of M. aeruginosa respond to varying eBeam doses and incubation periods post-irradiation. It is unknown at this time whether the DNA in cyanobacterial cells undergo multiple single and double strand breaks, and, therefore, DNA fragmentation studies should be performed to determine whether there are “hot-spots” in the DNA that are more susceptible to eBeam irradiation damage. Omic technologies such as transcriptomics, proteomics, and metabolomics should be employed to better understand how cyanobacterial cells respond to varying eBeam doses under differing experimental conditions. It is also important to determine whether inactivated cyanobacterial cells continue to produce toxins post irradiation. The potential for pre-formed toxins to be released from inactivated cells, as well as those potentially formed post-irradiation, must be understood. This information is imperative when designing an eBeam technology-based remediation strategy for the water industry.

Finally, there is a need to conduct research to enable the designing of an eBeam technology based treatment train for the drinking water industry. These studies should focus on demonstrating the degradation of the toxin and inactivation of the cyanobacterial cells in actual environmental surface water samples. Electron beam technology, if proven to be capable of inactivating the cyanobacterial cells and degrading the toxin molecule, can be used to remediate toxin containing drinking water in the treatment plants as well as detoxify drinking water treatment plant residuals containing cyanotoxins and toxin-producing cyanobacteria. The ability to treat such residuals can facilitate the disposable of hazardous wastes. Therefore, research should focus on identifying the doses required to attain specific toxin limits to facilitate disposal.

Overall, while there are promising results for the use of ionizing radiation technologies for the removal of MC-LR and M. aeruginosa in water, additional research is crucial in order to implement these technologies.