Removal of Microcystins from Drinking Water by Electrocoagulation: Upscaling, Challenges, and Prospects

Abstract

Microcystins (MCs) belong to a family of stable monocyclic heptapeptide compounds responsible for hazardous toxins in drinking water. Although several methods have been applied to remove MCs from drinking water (e.g., activated carbon filtration, ion exchange resins, high-pressure membranes, and electrochemistry), upscaling laboratory experiments to benefit municipal water treatment is still a major challenge. This chapter is a follow-up study designed to test three electrocoagulation (EC) techniques for decomposing MC by UV-ozone purification (laboratory), electrocoagulation (field unit), and coupled UV-ozone-electrocoagulation (municipal treatment). The chemistry and efficiency of the treatments were first examined followed by comparison with activated carbon filtration. Electrocoagulation outperformed activated carbon filtration by nearly 40%. When the laboratory treatments were evaluated at the municipal scale, effectiveness of the technique deteriorated by 10–20% because of UV pulse dissipation, vapor-ion plasma under-functioning, and limitations of polymer fiber filters. We confirmed previously published studies that pollutant coagulation and MC decomposition are affected by physicochemical factors such as radiation pulse density, electrical polarity, pH, and temperature dynamics. The results have relevant applications in wastewater treatment and chemical recycling.

1. Introduction

Cyanobacteria (also called cyanotoxins) in drinking water is a global concern because of their hazardous effects on human and animal health [1, 2, 3]. Microcystins (MCs) are a common source of cyanotoxins. MCs are produced by a variety of cyanobacteria including Microcystis spp, Anabaena spp, and Planktothrix 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. [4]. Microcystis aeruginosa is the most common species of cyanobacteria found in freshwaters around the globe and has been associated with a number of human, livestock, and wildlife poisoning [5, 6].

M. aeruginosa commonly produces microcystin-LR (MC-LR) (Figure 1) which is the most toxic and most prevalent of the over 100 identified variants [4, 5]. All MCs share a common structure including a cyclic heptapeptide containing three D-amino acids (alanine, glutamic acid, and methylaspartic 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) [7]. MC-LR contains leucine and arginine in the X and Z positions, respectively, and accounts for 99% of total harmful algal blooms [8]. Other less common variants include MCLA, MC-YR, MC-RR, MC-LF, and MC-LW. They are believed to have lower toxicity than MC-LR [6]. MC-LR’s biochemistry and toxicity are attributed to the ADDA moiety and its stereochemistry [9, 10]. Mechanistically, MC-LR targets hepatocytes in the liver and enters the cells by active transport aided by anion-transporting polypeptides [11, 12]. Next, the MC-LR binds strongly and irreversibly to serine or threonine protein phosphatases coded as PP1 and PP2A, which subsequently result in enzyme inhibition [13]. Given their importance in cell function and cell regulation, inhibition of PP1 and PP2A can result into 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 [12]. Acute toxicity can result in liver inflammation, hemorrhaging, and extensive hepatic bleeding. Death may occur due to liver failure at high or prolonged exposure.

Figure 1.

General structure of microcystin-LR.

MC-LRs are water-soluble and stable and demonstrate slow natural degradation (half-life = 10 weeks) in polluted water. The molecule is complex and heat-resistant making it toxic even after boiling. Although hard to remove by conventional water treatment, MC can rapidly degrade when exposed to UV radiation with wavelengths close to the absorption peak. Due to the presence of carboxyl, amino, and acylamino groups, MCs have been observed to ionize at temperatures above >40°C and in extreme acid-base media (pH <1.0 or pH >9.0) [9, 14, 15].

The distribution of MC in the US is a serious environmental health problem. Jenssen [16] has reported a wide range of MC concentration (12.5–225.6 μg/L) in multiple US communities. Environmental problems in the Wood County (West Virginia) and Mercer County (Ohio) closely reflect the national situation. Water quality data monitored between 2015 and 2019 by the EPA revealed that MC load in the Grand Lake St. Mary ranged from 0.0 to 79.7 μg/L, compared with the tolerable limit of 1.0 μg/L [17]. Similar data have been reported concerning the Ohio River Valley in West Virginia [16]. As mentioned before, UV exposure and electrocoagulation (EC) are useful methods for MC removal because of the capability to split their C–N bonds using electrical energy [15, 18, 19].

Recently, Folcik and Pillai [14] demonstrated the effectiveness of high-energy electron technology (advanced oxidation-reduction process) in degrading MC pollutants. The technology utilized accelerators to generate highly energetic electrons from regular electricity to create redox species to damage contaminants [20]. Similar examples of radiation technologies employed ⁶⁰Co gamma rays to inactivate MC multiplication [21, 22, 23]. Despite their effectiveness, these technologies are expensive and hi-tech and generally lack practical applications. Nevertheless, one of the techniques that are growing in popularity for MC decomposition is electrocoagulation [15].

Electrocoagulation (EC) employs the principles of electrochemistry for water treatment. It involves sacrificial corrosion of the electrodes (anode) to release active coagulant precursors (e.g., Al³⁺ or Fe²⁺) into solution. At the cathode, hydrogen gas evolves from electrolytic reactions. EC equipment can theoretically be scaled for any size and is not too difficult to operate. Recent technical improvements combined with a growing need for small-scale water treatment facilities have amplified interest in EC applications. Nonetheless, only a few studies have focused on the question of scale to demonstrate how laboratory filtration can be upgraded to municipal treatments. In addition, elucidating the key components that control EC production and MC removal efficiency is of paramount interest. Some of the factors that require illumination include current density, electrical polarity, and acid-base equilibria [24]. We hypothesize that a coupled UV-electrocoagulation process will completely remove MC from contaminated drinking water. We also predict that laboratory EC techniques are scalable to municipal purification cognizant that strong water treatment oxidizers like ozone are obtainable from the system’s vapor-ion plasma. The aim of this study is to (1) examine the operability and efficiency of cheap laboratory EC units for removing MC from drinking water, (2) test the scalability of laboratory EC filtration to municipal treatments, (3) evaluate the efficiency of the EC results against commercial water filtration (granulated activated carbon), and (4) examine the effects of radiation density, electrical polarity, pH, and temperature on the ionization of MC pollutants. The study will raise questions about electrocoagulation and industrial chemical recycling.

The chapter is structured in the following way. The first part reviews the literature on MC decomposition followed by description of the EC technique including the key components of the electrical units, electrodes activation, and reaction chemistry. The second section discusses the EC methodology followed by data generation, data analysis, and EC scalability. The final part examines the factors controlling EC physics, including radiation density, electrical polarity, pH, and temperature. The final section also discusses the economy of the new EC method.

2. Materials and methods

2.1 Equipment and raw materials

The basic raw material is surface and groundwater samples from the Mid-Ohio River Valley in Parkersburg (West Virginia) and untreated water from the Grand St. Mary’s Lake in Celina (Ohio). The illustration in Figure 2a represents the experimental design describing the general process of treating contaminated water using electrical energy. Figure 2b displays the laboratory-built UV-ozone water purification prototype, consisting of a 100-gallon plastic tank batch reactor fitted with an ionized nitrogen-oxygen (NI-OX) generator. The device is also fitted with a small fractional horsepower delivery compressor and 1-μm electron separation porous cellulose fiber water filter. Using basic engineering ideas, the unit was powered by a 110-V electrical source with the generator fastened to the tank cover and connected to a 1-μm polarized polymer filter suspended 10 inches above the inside base of the tank. The filter was connected to a fine bubble aeration diffuser using a half-inch poly tubing designed to eliminate debris, suspended solids, and microcystins pollutants. The principal component of the generator is a UV radiation lamp (λ = 155 nm) capable of splitting ambient gases (e.g., O₂ and N₂) into monoatomic-charged particles using ultraviolet ionizing energy and magnetic emission.

Figure 2.

(a) Schematic of the electrocoagulation purification system; (b) UV-ozone filtration unit; (c) Electrocoagulation unit; and (d) Coupled UV-ozone-electrocoagulation filtration system.

Figure 2c is a modified version of the prototype in Figure 2b designed to suit field conditions. It consists of a 400-gallon steel tank powered by a high-amperage (250 A), low-voltage generator (40 V) constructed to provide energy via switching polarity from direct current electric discharge. A characteristic component is 34 pairs of submerged anode and cathode crosslinked aluminum electrodes secured over a steel tank (Figure 2c). Crosslinked electrodes are the main reason for the (switching) reverse polarity. The coagulator works by establishing an intense electromagnetic field creating simultaneous oxidation-reduction reactions. An attached high-pressure pump was designed to channel polluted water over the metal plate contact areas. Treated water was pumped into a clean glass tank before samples are drawn for testing.

The third equipment (Figure 2d) is a coupled UV-EC (UV-ozone-EC) system, designed to mimic an industrial treatment system, with capacity to purify approximately 10,000 gallons of contaminated water within 12 h. Table 1 shows a summary of technical characteristics of the EC processing system.

	Scale	Raw Water Sample (Gallons)	Pump Horse-power	Energy from Generator	Minimum Time of Microcystin Decomposition (Minutes)	Maximum Time of Microcystin Decomposition (Minutes)
Water treatment method				Voltage (Volts)	Electrical current (Amperes)
UV-ozone ionization	Labo-ratory	90	1.5	9	80	20	50
Electro-coagulation	Field	400	2.5	24	160	10	30
UV-ozone-electro-coagulation	Muni-cipal	10,000	3.5	32	250	50	50
Activated carbon	Field	5	Not applicable	Not applicable	Not applicable	60	60

Table 1.

Characteristics of the water treatment techniques.

2.2 Experimental methods

Raw water and treated water samples were tested at the Industrial Chemical Laboratories, LLC (ICL) of Denver (CO). ICL is a specialized facility for testing chemical and biological pollutants in drinking water and wastewater including cyanobacteria. The samples were tested every 10 min for 90 min and analyzed using the Agilent 1100 tandem high-performance liquid chromatography-mass spectrometer (HPLC-MS). The glassware was thoroughly washed and rinsed with methanol and distilled water to prevent cross-contamination. The samples were first filtered with Whatman filter paper (1.2 μm) and chilled overnight at −20°C to dilute concentration of the pollutants. The filtrate was dissolved in 400 μL methanol and treated with a 2 mg/L sodium thiosulfate and acidified with trifluoroacetic acid (TFA, 0.1%, v/v), concentrated via solid phase cartridges (SuperClean LC-18, 3 mL tube), and eluted with 15 mL of 0.1% TFA in methanol. Aliquots of 20 μL were injected into system’s column (150 × 4.60 mm) at a flow rate of 1 mL/min at 30°C column temperature. The mobile phase consisted of H₂O plus 0.05% TFA and acetonitrile plus 0.05% TFA with a linear increase from 30 to 70% of the latter between 0 and 30 min. Chromatograms were recorded at 238 nm based on the literature. UV spectra and all chromatographic peaks were examined and compared to spectra standards of MC moieties. Peaks possessing the UV spectrum characteristic for MCs were quantified using a calibration curve. Unidentified peaks possessing the UV spectrum characteristic for MC but not matching the retention time of the standards were determined as MC-LR equivalents with a detection limit of 0.01 μg/L [25]. To understand seasonal variations, MC distribution was measured seven consecutive days in spring, summer, and fall and a regional mean calculated (Table 2).

	Range of water sampling depth (m)	Temperature (oC)	pH (-)	Turbi-dity (NTU)	Total dissolved Solids (mg/L)	Total organic carbon (mg/L)	Total nitrogen (mg/L)	Total phosphorus (μg/L)	Microcystins turbidity (μg/L)
Ohio River (Parkers-burg, WV)	0.0–1.5	23.3	8.58	0.41	121.4	3.8	0.62	27.23	112.6
Grand Lake St. Mary’s (Celina, OH)	0.0–1.5	22.9	8.55	0.67	126.7	3.1	6.88	301.91	147.5

Table 2.

Relationship between the microcystin and nutrient load.

The UV-ozone ionization reaction process was produced following the reactions below. Charged nitrogen particles were activated to release of free electrons (e⁻) to accelerate oxygen ionization:

$$N_2\frac{\mathit{UV}}{\mathit{\text{MagE}}}{N}^{\alpha +}-N^{\alpha -}\frac{e^{-}}{O_2\mathit{\text{ionized}}}\left(\text{free electron}\left[{\mathrm{e}}^{-}\right]\text{plus accelerating ionized}{\mathrm{O}}_2\right).$$E1

The oxygen radiation produced ozone vapor, ionized ozone, and superoxide ions and dissociated into more singlet oxygen (Eq. 2), resulting in a chain reaction of high-energy ionized oxygen in (Eq. 3):

$$O_2\frac{\mathit{UV}}{\mathit{\text{MagE}}}{O}_3\frac{\mathit{UV}}{\mathit{\text{MagE}}}\to {O_3}^{+}\leftrightarrow O_3^{-}\frac{\mathit{UV}}{\mathit{\text{MagE}}}\to O_2^{-}\left(\text{superoxide}\mathrm{ion}\right)$$E2

$$O-O^{-}-+{\mathit{xO}}^{-}-O^{-}\left(sin\mathit{\text{glet}}\right)-\frac{\mathit{UV}+\mathit{\text{electrons}}}{\mathit{\text{MagE}}}{}^{-}O-{}^{-}O-{}^{-}O\left(-O_x^{-}\right)-{}^{-}O-\left(\mathit{\text{chain}}\right)$$E3

The reaction of singlet oxygen (or chained ionized oxygen) with water was generated to produce high concentrations of hydrogen peroxide and/or hydroxide ions as saturated water produces excess peroxyl-reactive (oxidizing, disinfecting, and coagulating) ionized water in the subsequent reactions:

$${\mathit{xO}}_2^{-}+H_{2}O\frac{e^{-}}{\mathit{\text{stream}}}\to {\mathit{xH}}_2{O}_2\leftrightarrow {\mathit{xOH}}^{-}+{\mathit{xOH}}_2^{-}\left(\mathit{\text{scavenger}}\right)$$E4

$$O_2^{-}+H_{2}O\frac{e^{-}}{{\mathit{\text{MagE}}}_{\mathrm{\Delta }T}}\to H_2{O}_2$$E5

Thermal reaction of hydrogen peroxide and ozone was created to release free electrons and trioxidanes, superoxide ions, and peroxone Eqs. (6)–(8). The charged nitrogen and superoxide ions in aqueous solution were designed to produce additional free electrons, dinitrogen tetraoxide (nitroxyl ions), and hydroxide ions toxic to cyanobacteria:

$$H_2{O}_2+O_3\frac{e^{-}}{{\mathit{\text{MagE}}}_{\mathrm{\Delta }T}}\to H_2{O}_3+O_2^{-}$$E6

$$O_3+H_2{O}_2\frac{e^{-}}{\mathrm{\Delta }T}\to H_2{O}_5$$E7

$$N_2^{+}+3O_2^{-}+2H_{2}O\frac{e^{-}}{{\mathit{\text{MagE}}}_{\mathrm{\Delta }T}}{{}^{-}O}_2-N_2^{+}-O_2^{-}+4{\mathit{OH}}^{-}$$E8

The electrocoagulation reverse polarity reaction follows an electrolytic procedure [26]. The primary reactions at the anode and cathode are described in Eqs. (9)–(13):

$${\text{2H}}_2\mathrm{O}\left(l\right)\to {\mathrm{O}}_2\left(g\right)+{\text{4H}}^{+}\left(\mathit{aq}\right)+{\text{4e}}^{-}\text{Anode}$$E9

$${\text{4H}}_2\mathrm{O}\left(l\right)+{\text{4e}}^{-}\to {\text{2H}}_2\left(g\right)+{\text{4OH}}^{-}\left(\mathit{aq}\right)\text{Cathode}$$E10

$${\text{6H}}_2\mathrm{O}\left(l\right)\to {\text{2H}}_2\left(g\right)+{\mathrm{O}}_2\left(g\right)+{\text{4H}}^{+}\left(\mathit{aq}\right)+{\text{4OH}}^{-}\left(\mathit{aq}\right)\text{Overall}$$E11

$${\text{4H}}^{+}+{\text{4OH}}^{-}\left(\mathit{aq}\right)\to {\text{4H}}_2\mathrm{O}\left(l\right)$$E12

$${\text{2H}}_2\mathrm{O}\left(l\right)\to {\text{2H}}_2\left(g\right)+{\mathrm{O}}_2\left(g\right)$$E13

While reductants (free electrons) are released from the anode, oxidants and flocculation aggregates (e.g., H₂O₂, Al(OH)₃, Al₂O₃) are generated at the cathode as shown in Eqs. (14)–(18):

$${\mathit{Xe}}^{-}+H_{2}O\stackrel{\mathit{\text{Metal Electrodes}}}{\to }{H}_2\uparrow +O_2\uparrow \stackrel{\mathit{\text{Electrons}}}{\to }{O}^{-}-O^{-}-O_x^{-}\left(\mathit{\text{Oxidation}}\right)$$E14

$$\stackrel{\mathit{\text{Electrons}}}{\to }{\mathit{OH}}^{-}+H_2{O}_2$$E15

$$\stackrel{\mathit{\text{Electrons}}}{\to }\mathit{Al}{\left(\mathit{OH}\right)}_3\downarrow$$E16

$$\stackrel{\mathit{\text{Electrons}}}{\to }{\mathit{Al}}_2{O}_3\downarrow$$E17

$$\stackrel{\mathit{\text{Electrons}}}{\to }\mathit{\text{Electrophilic}}+\mathit{\text{Nucleophilic cleavage of}}C-F\mathit{\text{bonds}}$$E18

The pollutant removal efficiency (% r) was calculated using Eq. (19) as follows:

$$\%r=\left(\frac{C0-\mathit{Ct}}{C0}\right)x100\%$$E19

where C₀ is the initial concentration of pollutant and C_t is the concentration of pollutant at time t.

To initialize the EC process and augment flocculent formation, about 80 g of potassium aluminum sulfate dodecahydrate (KAl₂(SO₄)₂.12H₂O) (potash alum) solution was used both as an electrolyte and coagulant following previous experiments by Johnson [26]. Finally, treated water samples were compared with data from a commercial gravity block ionic adsorption unit fitted with granular activated carbon filters and coated by silver-impregnated ceramic outer shells. The experimental results are discussed in the subsequent section.

3. Results and discussion

3.1 Electrocoagulation and scalability

The results in Figure 3 displays microcystin (MC) response to laboratory UV exposure compared with field and municipal electrocoagulation (EC). To determine the EC production efficiency, MC filtration data were compared with the WHO’s maximum contaminant level of 1.0 μg/L. While the laboratory and field experiments decomposed MC within 10–20 min, the municipal system disrupted MC bonds after 50 min. Multiple reasons can explain this. The first one is a technical challenge. As expected, building and testing the 10,000-gallon reactor (Table 1) was more arduous than the 90-gallon reactor. The installation of high-intensity UV lamps to generate optimal radiation density in the larger reactor was particularly challenging. Another difficulty was how to generate maximum turbulence to aerate and circulate radiation. Although this was improved using bubble diffusers, predicting diffuser size was still time consuming, requiring several iterations. The reaction delays were also attributable to differences in surface energy interactions between radiation pulse and pollutant substrates. A recent study by Cavitt et al. [27] has shown that molecular bond disruption in aqueous media is controlled by many thermodynamic factors such as reaction rates, solvent volume, acid-base equilibria, and interfacial alignment of reactants versus products. Given similar temperature doses, reaction rates were better favored in the laboratory (90-gallon) reactor than its municipal (10,000-gallon) counterpart. The general conclusion is that MC bond disruption is easier in smaller reaction tanks than larger ones.

Figure 3.

Comparison of microcystin removal techniques.

The aforementioned result notwithstanding the results in Figure 3 shows a slight deviation between UV irradiation (20 min) and EC (electrolytic) treatments (10 min). Notice that UV irradiation is closely related to physical factors such as UV lamp size, pulse intensity, and radiation diffusion (15, 18), while EC is controlled by direct electrical vibration against C–N bonds. The aforementioned, therefore, is a reasonable explanation for the observed discrepancy. Another reason for the reaction delay is ozone deficiency possibly resulting from coupling glitches between the system’s ionized nitrogen-oxygen (NI–OX) generator and its compressor (Figure 2b). This matter will be further investigated in subsequent studies. It is worth noting, however, that all the prototypes (Figure 2) decomposed MC molecules reasonably well. Notice for instance, that the municipal EC unit destroyed MC by approximately 80% within the first 10 min, while the coupled UV-ozone treatment was even better at 95%. Studies such as Wolfe et al. [28], Langlais et al. [29], and Folcik et al. [14] have reported polar bond disruption from electron bombardment, free radical attack, and ozone and peroxone toxicity. Peroxone is a mixture between ozone and hydrogen peroxide Eqs. (1)–(8). The theoretical basis is that heavy oxidizing agents (e.g., peroxides, trioxidanes, and peroxones) can break down functional C–N bonds of microcystins [30, 31]. Previously, He et al. [32] reported the destruction of cyanobacteria using hydroxyl-free radicals. In addition, studies such as Yoo et al. [33] and Lui et al. [34] have demonstrated how low doses of peroxyl and nitroxyl ions can disrupt chemical bonds in molecular compounds. Results from this study quite closely reflect conclusions by previous researchers.

To validate MC decomposition data in Figure 3, UV treatments were matched against granular activated-carbon filtration data. The carbon filtration was from a commercial source and readily available. The results are displayed in Figure 4a. Two important things are observable from the outcomes. While the UV filtration disrupted C–N bonds within 20 min, the activated carbon produced treatments after 30 min. Still, the UV filtration outperformed the commercial granular carbon filtration by nearly 40%. This was expected knowing that UV radiation is more energetic in splitting C–N bonds. To predict the rate of MC removal, the data were subjected to a crude regression analysis. The curve followed a polynomial decay in the form of y = 8.394x²–88.635x + 221.97; R² = 0.9399 (Figure 4b) confirming the robustness of UV radiation in removing MC pollutants within 1 h. This knowledge is relevant and applicable to wastewater treatment and chemical recycling.

Figure 4.

(a) Comparison between UV radiation and granular activated-carbon removal of microcystins; and (b) Regression curve of the UV-ozone water treatment.

3.2 Principal component analysis

This section discusses electrocoagulation (EC) principal components that control MC decomposition. The parameters below were considered important in the published literature [15]: (a) voltage (as proxy data for radiation density), (b) pH (acid-base equilibria), (c) electrical polarity (reverse polarity), and (d) temperature. The parametric data were derived using Eq. (19). The results are displayed in Figure 5. The municipal EC results were evaluated using the published data by Miao et al. [31] and further verified [35]. Two important points are observable here. First, MC removal increased with increasing radiation density. Second, there was a difference in optimal voltage density coincidental with maximum MC decomposition. While the field reactor completely removed MC (100% decomposition) at 24 V, the municipal reactor performed maximally at 32–40 V showing 95% pollutant removal (Figure 5a). The difference in energy dosage was attributed to the sheer size of the municipal reactor, which in turn resulted from generator adjustments to solve solvent turbulence and flocculent formation deficiencies.

Figure 5.

Factors controlling electrocoagulation efficiency in decomposing microcystin (a) radiation density; (b) bond polarity; (c) pH; and (d) temperature.

As expected, no large differences were observed between the field and municipal reactions in terms of acid-base equilibria and thermodynamics (Figure 5b and d). The most significant conclusions are that (1) pH and temperature elevations are more favorable to MC decomposition; and (2) optimal pH for pollutant removal lies in the alkaline range with pH > 8.00. While the pH data strongly agreed with recent findings by Folcik and Pillai [14], it was strikingly contrary to previous conclusions by Bao et al. [16] whose studies on C–F decomposition was more productive in acidic media. It is possible to explain our findings in two ways. First, the coagulation “seeding agent” (i.e., potash alum (KAl₂(SO₄)₂.12H₂O) may have contributed to pH elevation. Secondly, the production of metal hydroxides such as Al(OH)₃ and Fe(OH)₃ from sacrificial anodes (aluminum and iron metals electrodes in the electrolytic cell (Figure 2)) may have produced more alkaline conditions. Thermodynamic effects on MC decomposition are well researched including published articles by Folcik and Pillai [14], Folcik et al. [36], and Wang et al. [37]. The conclusion here is that MC bonds are significantly disrupted at temperatures beyond 40°C. The results in Figure 5d quite closely matched some of the aforementioned findings. In this study, however, 30°C was observed as the starting point of MC decomposition with maximum disruptions encountered between 70 and 90°C. The difference may largely be due to high generator amperage (250 A) employed (Section 2.1).

Another important EC factor is electrical polarity. Previous studies such as Triantis et al. [38] and Gajda et al. [39] have reported limitations of conventional single-anode polarity in EC production. For this reason, we experimented a more robust switching polarity procedure using crosslinked aluminum electrodes with its energy from direct current electrical discharge. Using a trial-and-error optimization approach, the reactor was “trained” to switch electrical current bombardments between the anode and cathode electrodes. The data in Figure 5c show that MC removal was maximum (100% removal) at every 5 s. The question is: Why is the switching polarity so important? The answer relates to two important things. First, optimizing the system saved time, power, and ultimately, cost. Second, the switching polarity ensured that the C–N bonds were attacked at both the anode and the cathode, contrary to conventional one-way electrical bond splitting. The dual attack against C–N bonds is a major reason for effective EC production. Notice, however, that the response from the large-size municipal reactor was inferior compared with its field counterpart. The discrepancy is still not easy to explain. However, operational problems such as electrode size determination for maximum flocculent distribution may be responsible for deviations. This is another topic worth investigating in future studies.

How does the new EC equipment compare with conventional community water treatment in terms of cost economics? To answer this question, the municipal prototype was “starved” of ozone and UV radiation, while extending treatments beyond 90 min. The goal was to examine whether the EC system would provide cheaper filtration compared with conventional treatments in the study area. The results are displayed in Table 3. The data show that while groundwater MC treatment was unimportant in West Virginia, the importance of surface water treatment was without question. The heavy pollution associated with both the Ohio River and Grand St. Mary’s Lake (Table 2) is noticed. The data in Table 3 show that the EC procedure was much cheaper than conventional membrane filtration or chemical disinfection. Specific to MC removal, the EC method was predicted to be nearly 800 times cheaper than conventional treatments at the Celina plant. On the basis of this study, incorporating EC methods at conventional treatment plants has potential to both improve water treatment chemistry and save cost.

	Celina (Ohio) Grand Lake St. Mary’s	Parkersburg (West Virginia)
Source of Drinking Water	Lake water	Groundwater (Ohio River Valley)
Scale of Water Treatment	Municipal	Municipal
Rate of Water Treatment	1000 gallons/minute	265 gallons/minute
Cost of Water Treatment	$3.66/1000 gallons/minute	$1.84/1000 gallons/minute
Cost of Microcystin Removal using Conventional Techniques (Aeration, bio-digestion & membrane filtration)	$0.37/1000 gallons/minute	$0.00/1000 gallons/minute*
Cost of Microcystin Removal by Electrocoagulation	$0.04/1000 gallons/minute	$0.04/1000 gallons/minute

Table 3.

Costs comparison between conventional water treatment and new electrocoagulation.

*

Parkersburg groundwater treatment is achieved by activated carbon filtration.

3.3 Further studies

This study has confirmed published reports that advanced EC methods are effective in removing MC pollutants from drinking water. However, some key topics remain to be investigated including (a) chemical elucidation of decomposed MC fragments, (b) scalability of UV lamps, ionized nitrogen-oxygen (NI-OX) generators and EC electrodes, (c) technical operability and cost assessments, and (d) applicability of the technique for wastewater treatment and chemical recycling. Some of these topics will be addressed in subsequent studies.

4. Conclusion

The removal of microscopic pollutants in drinking by electrocoagulation (EC) is becoming increasing popular because of the use of radiation energy in decomposing molecules that contain polar bonds including C–F and C–N bonds. The purpose of this study was to discuss three EC techniques for removing microcystins (MC) in contaminated drinking water at the Celina (OH) and Parkersburg (WV) treatment plants and to compare their effectiveness at the laboratory, field, and municipal scales. While the laboratory and field experiments employed UV-ozone and electrolytic cell filtration techniques, respectively, the municipal experiment applied a coupled UV-ozone and EC technique. To validate the effectiveness of the methods, the EC results were evaluated against a commercially available granular activated carbon filtration unit. The EC technique outperformed the activated carbon filtration by more than 40%. When the laboratory treatments were upscaled and tested at a municipal level, effectiveness of the technique declined by nearly 10–20% because of pulse dissipation from UV lamps, vapor-ion plasma underactivity, and limitation of membrane filters. We confirmed previously published studies that pollutant coagulation and MC decomposition were affected by physical factors such as radiation density, reverse electrical polarity, pH, and temperature. These results have other applications in industrial wastewater treatment and chemical recycling.

Acknowledgments

The authors are very grateful to the Vienna City and Mercer County Councils for funding the chemical data analyses. We are also grateful to the biochemistry majors of Ohio Valley University for assisting coagulator test runs and data collection. Finally, we appreciate Dr. Matt Vergne of Lipscomb University, Nashville, Tennessee, for his critical review of the original manuscript.

Conflict of interest

The authors declare no competing financial interests.