Abstract

The use of constructed wetlands as a wastewater treatment system is a feasible solution for rural areas. However, these systems do not efficiently eliminate pathogenic microorganisms. Therefore, it is necessary to implement disinfection systems such as ultraviolet (UV) disinfection systems in constructed wetlands. To evaluate the behavior of a UV system, a pilot system of artificial wetlands connected to one such disinfection system was operated. The results show that when the total suspended solids (TSS) of the influent (already treated by the system of constructed wetlands) reached values of 26.7 mg/L, a reduction of 2.03 uLog in fecal coliforms was obtained. However, when the TSS increased to 34.7 mg/L, the reduction was only 0.33 uLog. In addition to the influence of the TSS on the fecal coliform reduction efficiency, there is a direct relationship between the transmittance and the sizes of the particles present in the influent. After UV treatment, the microorganisms showed a peak in photoreactivation of 27.8% at 4 h after irradiation with visible radiation, while under conditions of darkness, no reactivation was observed.

INTRODUCTION

Constructed wetlands (CW) are a solution for the treatment of wastewater in rural areas because CW are simple to operate and have low maintenance costs. Reportedly, there are more than 50,000 wastewater treatment plants (WWTPs) in Europe and more than 10,000 in North America based on CW (Yan & Xu 2014). Vera et al. (2011) found that treatment of wastewater from small communities (<2,000 people) with CW yielded reductions of 5-day biological oxygen demand (BOD5) between 78.4 and 96%, total suspended solids (TSS) between 65.2 and 88.3%, total nitrogen (TN) between 48.1 and 65.4% and total phosphorus (TP) between 39.4 and 58.4%. However, these systems do not efficiently inactivate pathogenic microorganisms in wastewater. López et al. (2019) showed that the wastewater disinfection by CW with horizontal sub-surface flow (HSSF) yields reductions between 0.24 and 3.63 Log colony forming units (CFU)/100 mL for fecal coliforms and between 0.46 and 2.84 Log plaque forming unit (PFU)/100 mL for coliphages. Consequently, after treatment by CW, the treated effluents contain between 1.65 and 5.38 Log CFU/100 mL of fecal coliforms and 0.10 and 1.41 Log PFU/100 mL of coliphages. Thus, CW systems must have a disinfection unit, such as an ultraviolet (UV) system, because these systems are easy to operate and secure and are economically viable (Chatterley & Linden 2010). Azaizeh et al. (2013) showed that a CW system achieves only a 1.5 uLog reduction in fecal coliforms. However, if the CW is coupled to a UV system, a reduction of 9.1 uLog was possible, where 1 log = 90% disinfection, 2 log = 99% disinfection and so on.

UV and ultrasound systems are defined as physical agents that have the distinction of not generating disinfection byproducts that may affect the health of humans or that of biotic organisms present in bodies of water (Neumann et al. 2017; Vázquez-López et al. 2019). Also, advanced oxidation processes are processes that appear to be more effective in deactivating pathogens (Vidal et al. 2004; Yeber et al. 2009). Unlike chlorine disinfection, UV more effectively inactivates Giardia lamblia and Cryptosporidium parvum (Betancourt & Rose 2004; Leiva et al. 2019).

The transmission of UV radiation is mainly affected by suspended solids or the turbidity of the influent to be treated. The solids in the suspension absorb light, contributing to shading effects, protecting microorganisms from UV exposure (Emerick et al. 1999) and affecting the transmittance of the water to be treated. Particles larger than 7 μm present in the influent to be treated can affect the disinfection process much more than particles smaller than 7 μm by forming particle–microorganism associations (Jolis et al. 2001). This particle–microorganism association is strongly regulated by the size and zeta potential of the particles (Chahal et al. 2016). The zeta potential provides information on the electrostatic repulsive forces that may exist between the particles and the microorganisms present, aiding understanding of the possible formation of biofilms that would directly affect the functioning of the UV system (Bhattacharjee 2016).

Among the disadvantages of using these UV systems it has been found that certain microorganisms have been able to develop DNA repair mechanisms. This is in order to deal with the damage caused by UV radiation when low doses are applied in the UV-C spectrum ranges (Zhou & Smith 2002). This photoreactivation that occurs in the presence of visible light has been found, where the photolyase enzymes act by reversing the damage caused by UV light, and reactivation in the dark where the damaged DNA is replaced by new nucleotides (Friedberg 2016; Li et al. 2017). Therefore, it is essential to control the microorganism behaviors to ensure proper disinfection by UV systems.

Considering the above, the objective of this study was to evaluate the behavior of a UV system for the disinfection of an effluent from a pilot system of wetlands and to determine the rates of reactivation of pathogenic microorganisms to evaluate the actual efficiency of the UV system.

MATERIALS AND METHODS

Pilot plant and sampling strategy

The study was carried out in a pilot plant that has been in operation for 8 years (since 2011) and is located in Hualqui (36°59′26.93″S and 72°56′47.23″W), in the Biobío region of Chile. In this location, marked seasonal trends were observed regarding precipitation, with an average of 71 mm of rain during the summer season, while for autumn, winter and spring it was 226, 320 and 112 mm, respectively. The influent of this plant is wastewater that has been subjected to a primary treatment in a WWTP that serves a rural community of 20,000 inhabitants.

This pilot plant consists of six HSSFs, which are called wetland cells, with a surface area of 4.5 m2 each and a water depth of 0.4 m. Phragmites australis, Schoenoplectus californicus, Zantesdeschia aethiopica and Cyperus papyrus were planted in these cells. The support material is gravel sized 3/4″ to 1″ (1″ = 2.54 cm), and the hydraulic retention time varies between 3 and 7 days (López et al. 2015). A UV radiation system was added to this pilot plant as an additional disinfection system. The influent used in this study was a mixture of the water treated through the six wetland cells, and the effluent was the output of the UV system.

The system used was the disinfection reactor model HO UV30-187OC (Hidro-UV), composed of a UV lamp with a disinfection power of 30 mJ/cm2. The mercury lamp corresponded to a type of high-output lamp of low pressure and high intensity, with an input power of 87 W and an output of 28 W. The protective covers of this lamp are composed of 99% SiO2, and the UV system flow is gravitational, with an input and output connection of 1″. The UV disinfection equipment was operated at 0.4 m3/h because, at the time of installation of this system, the transmittance analysis yielded a value of 20%. Monthly maintenance was performed by cleaning the quartz tube that covers the UV lamp according to the procedure indicated by the supplier.

Samples were taken in each season of the year in which the study was conducted, corresponding to winter, spring and summer. A sample of the influent and another of the effluent was collected, for which the physicochemical characteristics were measured and the disinfection power was evaluated by the analysis of the fecal coliforms and total and somatic coliphages. The samples were collected in sterile bottles and transported to the laboratories of the Environmental Engineering and Biotechnology Group (Grupo de Ingeniería y Biotecnología Ambiental – GIBA) of the University of Concepción, where the samples were analyzed 2 hours after collection. Figure 1 shows the arrangement of the pilot plant, the UV system and the analyses performed.

Figure 1

Schematic diagram of pilot-scale wastewater treatment plant using a UV disinfection system with sampling points, and the physicochemical and microbiological analyses.

Figure 1

Schematic diagram of pilot-scale wastewater treatment plant using a UV disinfection system with sampling points, and the physicochemical and microbiological analyses.

Analytical methods

Characterization of the influent and effluent

The in situ parameters (temperature, pH, electric conductivity, turbidity, and dissolved oxygen (DO)) were measured by OAKTON portable multiparameter equipment (PC650-480485), a portable OAKTON (T-100) waterproof turbidimeter and a portable oximeter (oxi 330i/set Hanna HI 9146-04). The physical-chemical parameters (TSS, volatile suspended solids (VSS), TN, ammoniacal nitrogen (NH4+-N), TP, phosphate (PO4−3-P), chemical oxygen demand (COD), and BOD5) were measured according to the protocols described in standard methods (APHA 2005). For the TN and TP, specific kits from the Merck Spectroquant line were used. The microbiological parameters (fecal coliforms (FC), total coliforms (TC) and somatic coliphages) were determined by the MPN technique, in which there is a presumptive test and another confirmatory test for the presence of the bacterial group as indicated in standard method 9221-TC (APHA 2005). The amounts of somatic coliphages were determined by the ‘single agar layer’ procedure using Escherichia coli as the host strain. This methodology is based on the principles of protocols 9224 B and 9224 E of the standard methods (APHA 2005).

In addition, the transmittance, particle size and zeta potential were measured in the samples obtained in the spring and summer seasons. For the transmittance analysis, a UV-Vis Thermo Spectronic spectrophotometer (Genesis 10 UV) at 254 nm was used. For the particle size, Microtrac FLEX equipment was used, and for the analysis of the zeta potential, ParticleMetrix equipment from Stabino was used.

Reactivation analysis

To measure the reactivation of the microorganisms, a Petri dish was used to which 20 mL of the sample of water treated by the UV system in the pilot plant was added. The visible lamp model OSRAM L 15 W/840 of LUMILUX was located 30 cm away from the liquid surface of the sample.

The photoreactivation period was 24 hour in total, and the samples were taken at intervals of 0, 2, 4, 6 and 24 hours. Another set of experiments was carried out in darkness under the same conditions as those of the samples exposed to light. In addition, a control was carried out in which the photoreactivation of the samples was studied without prior exposure to UV radiation.

The reactivation was expressed as a function of the survival ratio using Equation (1):  
formula
(12)
where S is the survival ratio at time ‘t’, N0 is the number of microorganisms before UV irradiation (MPN/100 mL) and Nt is the number of microorganisms after photoreactivation for a period of time ‘t’ (MPN/100 mL). This survival ratio gives the final effect of UV inactivation when reactivation is considered (Li et al. 2017).

Statistical analysis

For the influent and UV effluent data, only the microbiological parameters were compared. To statistically analyze the results obtained, the statistical program InfoStat (Di Rienzo et al. 2011) was used. The data were subjected to normality tests using the Shapiro–Wilk test to determine the appropriate tests to be applied. To compare the effluent and the influent, the normally distributed data were analyzed using a paired t-test, and data not normally distributed were analyzed using a Wilcoxon test. In all statistical tests, a significance level of α = 0.05 was used.

RESULTS AND DISCUSSION

Characteristics of the effluent of the constructed wetland

Figure 2 shows the behavior of different parameters measured in situ in the effluent of the HSSF where pH values between 6.6 and 8.1 are found in different seasons of the year. The temperatures are mainly a function of the characteristics of each station studied, being specific to the area where the pilot system is located and varying between 11.6 ± 2.1 °C in the winter and 20.3 ± 1.7 °C in the summer. For the oxidation reduction potential (ORP), average values of −14.0 ± 200.2 mV are found in the winter, corresponding to the anaerobic and reduction values of the type of CW studied (Leiva et al. 2019). In spring and summer, the ORP values average −6.5 ± 167.4 mV and 198.7 ± 22.0 mV, respectively, presenting mostly aerobic characteristics. This behavior may be due to the conditions in which the samples were taken because the samples may have been exposed to the atmosphere, which can change the properties of the samples.

Figure 2

Characterization of in situ parameters of the HSSF effluent for (a) temperature, (b) pH, (c) dissolved oxygen and (d) ORP. For winter, n = 3; for spring, n = 4 and for summer, n = 3.

Figure 2

Characterization of in situ parameters of the HSSF effluent for (a) temperature, (b) pH, (c) dissolved oxygen and (d) ORP. For winter, n = 3; for spring, n = 4 and for summer, n = 3.

The DO levels in effluents from CW are usually lower than 2.0 mg/L (Masbough et al. 2005). In the current study, the DO contents for winter, summer, and spring are 95%, 10%, and 25% above 2.0 mg/L, respectively.

Table 1 shows the physicochemical parameters of the HSSF effluent. The elimination of nutrients appears to be intimately related to the temperature fluctuations of the different seasons. In the summer, 41.3 mg/L of NH4+-N is detected at 22.0 °C, while in the winter, the highest value of 71.4 mg/L of NH4+-N is obtained at 10.1 °C. In this sense, increasing temperature promotes the oxidization of NH4+ by nitrifying microorganisms, and an anaerobic environment favors denitrifying organisms that can reduce and eliminate these compounds in the liquid phase into the atmosphere (Vymazal 2007).

Table 1

Characterization of physical-chemical parameters of the HSSF effluent

ParameterConcentration (mg/L)
Winter
Spring
Summer
AverageRangeAverageRangeAverageRange
TSS 26.7 ± 8.4 17.0–32.4 31.8 ± 15.9 19.5–55.0 34.7 ± 23.5 7.2–64.0 
VSS 28.3 ± 1.9 13.9–29.6 23.7 ± 6.9 15.0–31.6 26.6 ± 17.1 6.8–48.5 
CODt 118.6 ± 32.9 82.4–146.6 161.1 ± 41.2 124.3–215.5 161.8 ± 76.1 115.2–249.6 
CODS 97.3 ± 4.3 53.4–100.4 124.4 ± 25.6 89.0–150.2 104.5 ± 31.9 75.2–138.5 
BOD5 76.6 ± 7.9 70.5–85.5 95.4 ± 24.3 74.4–127.5 83.6 ± 22.8 67.5–99.7 
TN 72.0 ± 8.5 66.0–85.0 66.8 ± 3.9 62.0–70.0 78.7 ± 6.8 71.0–84.0 
NH4+-N 53.6 ± 18.2 35.0–71.4 34.1 ± 7.5 26.2–44.4 31.7 ± 19.0 9.8–43.9 
TP 13.2 ± 1.8 9.0–14.4 13.2 ± 1.0 12.0–14.3 13.3 ± 1.1 12.6–14.6 
PO4−3-P 7.7 ± 0.3 7.4–8.0 12.0 ± 2.4 8.5–13.8 12.6 ± 1.1 11.4–13.7 
ParameterConcentration (mg/L)
Winter
Spring
Summer
AverageRangeAverageRangeAverageRange
TSS 26.7 ± 8.4 17.0–32.4 31.8 ± 15.9 19.5–55.0 34.7 ± 23.5 7.2–64.0 
VSS 28.3 ± 1.9 13.9–29.6 23.7 ± 6.9 15.0–31.6 26.6 ± 17.1 6.8–48.5 
CODt 118.6 ± 32.9 82.4–146.6 161.1 ± 41.2 124.3–215.5 161.8 ± 76.1 115.2–249.6 
CODS 97.3 ± 4.3 53.4–100.4 124.4 ± 25.6 89.0–150.2 104.5 ± 31.9 75.2–138.5 
BOD5 76.6 ± 7.9 70.5–85.5 95.4 ± 24.3 74.4–127.5 83.6 ± 22.8 67.5–99.7 
TN 72.0 ± 8.5 66.0–85.0 66.8 ± 3.9 62.0–70.0 78.7 ± 6.8 71.0–84.0 
NH4+-N 53.6 ± 18.2 35.0–71.4 34.1 ± 7.5 26.2–44.4 31.7 ± 19.0 9.8–43.9 
TP 13.2 ± 1.8 9.0–14.4 13.2 ± 1.0 12.0–14.3 13.3 ± 1.1 12.6–14.6 
PO4−3-P 7.7 ± 0.3 7.4–8.0 12.0 ± 2.4 8.5–13.8 12.6 ± 1.1 11.4–13.7 

All values are expressed as mean ± standard deviation and range of minimum and maximum. TSS: total suspended solids; VSS: volatile suspended solids; CODt: total chemical oxygen demand; CODS: soluble chemical oxygen demand; BOD5: 5-day biological oxygen demand; TN: total nitrogen; NH4+-N: ammoniacal nitrogen; TP: total phosphorus; PO4−3-P: phosphate. For winter, n = 3; for spring, n = 4 and for summer, n = 3.

With respect to TSS results and turbidity, an inverse relationship is presented as shown in Figure 3. In winter, values for TSS of 26.7 mg/L were found while turbidity was 45.4 NTU, while in summer the TSS increased to 34.7 mg/L and turbidity decreased to 21.7 NTU. This finding could be considered unexpected, because it is assumed that the TSS is responsible for the turbidity and color of the water. However, Bilotta & Brazier (2008) mention that turbidity responds to factors other than the concentration of TSS, so there is not always a direct relationship between these two parameters. Similarly, Bilotta & Brazier (2008) report that turbidity measurements are influenced by various characteristics, such as the particle size, the shape of the TSS, the presence of phytoplankton and the presence of dissolved mineral substances.

Figure 3

Relationship between the amount of TSS (bar chart) and turbidity (♦) in the different stations studied. For winter, n = 3; for spring, n = 4 and for summer, n = 3.

Figure 3

Relationship between the amount of TSS (bar chart) and turbidity (♦) in the different stations studied. For winter, n = 3; for spring, n = 4 and for summer, n = 3.

In addition, solutions with equal concentrations of suspended solids but of different compositions do not disperse the same amount of light, confirming that turbidity may not have a direct relationship with the concentration of TSS, raising several concerns (Pavanelli & Pagliarani 2002; Bertrand-Krajewski et al. 2010; Hannouche et al. 2011).

Figure 4 shows that there is a direct relationship between the amount of organic matter (in this case represented as CODt) and the amount of TSS. In winter, TSS values of 26.7 ± 8.4 mg/L and CODt values of 118.6 ± 32.9 mg/L were obtained, while in summer, an average TSS of 34.7 ± 23.5 mg/L was obtained with a CODt of 161.8 ± 76.1 mg/L. This effect could be associated with the higher generation rate of biofilms in warm seasons. The formation of these biofilms induces an increase in the amount of TSS as the amount organic matter increases. This effect, together with an average temperature of 20.3 ± 1.7 °C in summer and 11.6 ± 2.1 °C in winter, generates 23% higher TSS in summer compared to that in winter.

Figure 4

Relationship between the amount of organic matter (bar chart) and the amount of TSS (▪) in the different stations studied. For winter, n = 3; for spring, n = 4 and for summer, n = 3.

Figure 4

Relationship between the amount of organic matter (bar chart) and the amount of TSS (▪) in the different stations studied. For winter, n = 3; for spring, n = 4 and for summer, n = 3.

Analysis of transmittance, zeta potential and particle size

Figure 5(a) shows the results obtained from the transmittance analysis. The values measured for the spring season are 23.9% for the influent and 27.8% for the effluent, while for the summer, values of 29.0% are obtained for the influent and 29.3% for the effluent. These values are likely associated with the turbidity levels of 22.4 NTU in spring and 21.7 NTU in summer. For UV radiation to function effectively, water can have an acceptable transmittance between 35% and 65% so that the UV system has the desired impact on microorganisms (Das 2001).

Figure 5

Results obtained from analysis of (a) transmittance, (b) zeta potential and (c) size of particles for the influent (▪) and effluent (□) in spring and summer. For spring, n = 3, and for summer, n = 3.

Figure 5

Results obtained from analysis of (a) transmittance, (b) zeta potential and (c) size of particles for the influent (▪) and effluent (□) in spring and summer. For spring, n = 3, and for summer, n = 3.

According to the recommendations of the supplier (Hidro-UV), the transmittance values should be higher than 60% for more effective elimination of pathogens, while the optimum is 80% to ensure maximum disinfection effectiveness. These recommendations are not met in this study, where the highest transmittance is 33.7% in summer. These results suggest that the transmittance affects the optimum disinfection.

Regarding particle size, Emerick et al. (1999) mention that bacteria are generally associated with particles larger than 10 μm, while according to Mattle & Kohn (2012), viruses can associate with particles smaller than 2 μm. Figure 5(b) shows an average particle size of 22.4 μm for the samples analyzed in summer, with 95% of the particles present sized 54.4 μm.

Microbial binding to the surfaces of the particles is driven by electrostatic interactions, van der Waals forces, hydrophobicity, surface tension and surface roughness (Hassard et al. 2016). In this sense, the zeta potential represents the surface charges of the particles present. In the samples analyzed, particles with negative surface charges are found, as shown in Figure 5(c). In the spring, average values of −15.3 mV and −15.6 mV are obtained for influent and effluent, respectively, and in the summer, values of −15.5 and −15.3 mV are obtained for the influent and effluent, respectively. These negative values suggest that the particles have a greater repulsive energy barrier than particles with zeta potential values close to 0 mV.

Disinfection by UV system

Table 2 shows the TC, FC and somatic coliphage results obtained after UV system disinfection. Considering all the TC removal results obtained per station, the lowest TC concentration is obtained in winter, with 1.6 · 104 MPN/100 mL at the output of the UV system. For the elimination of FC obtained per station, the lowest concentration is obtained in winter, with 1.6 · 104 MPN/100 mL at the exit of the UV system. For the somatic coliphages, the variability in the average reduction was not marked by seasonality, given that the highest average was found in winter and then in spring. In this case, the lowest concentration of somatic coliphages was obtained in summer at 1.0 · 102 CFU/100 mL.

Table 2

Microbiological characterization by total and fecal coliforms of the HSSF effluent seasonally

Concentrations
Total coliformFecal coliformSomatic coliphages
MPN/100 mLMPN/100 mLCFU/100 mL
Winter Average 5.3 · 106 ± 1.4 · 105 1.6 · 106 ± 7.1 · 104 2.0 · 104 ± 4.5 · 103 
Range 5.2 · 106–5.4 · 106 1.5 · 106–1.7 · 106 1.7 · 104–2.0 · 104 
Spring Average 7.1 · 106 ± 5.2 · 106 4.1 · 106 ± 3.4 · 106 7.4 · 103 ± 7.1 · 102 
Range 3.3 · 106–1.3 · 107 1.1 · 106–7.8 · 106 6.9 · 103–7.9 · 103 
Summer Average 1.2 · 107 ± 1.5 · 107 9.2 · 106 ± 1.7 · 107 1.8 · 104 ± 2.2 · 104 
Range 4.9 · 106–3.5 · 107 4.5 · 105–3.5 · 107 1.3 · 103–4.8 · 104 
Concentrations
Total coliformFecal coliformSomatic coliphages
MPN/100 mLMPN/100 mLCFU/100 mL
Winter Average 5.3 · 106 ± 1.4 · 105 1.6 · 106 ± 7.1 · 104 2.0 · 104 ± 4.5 · 103 
Range 5.2 · 106–5.4 · 106 1.5 · 106–1.7 · 106 1.7 · 104–2.0 · 104 
Spring Average 7.1 · 106 ± 5.2 · 106 4.1 · 106 ± 3.4 · 106 7.4 · 103 ± 7.1 · 102 
Range 3.3 · 106–1.3 · 107 1.1 · 106–7.8 · 106 6.9 · 103–7.9 · 103 
Summer Average 1.2 · 107 ± 1.5 · 107 9.2 · 106 ± 1.7 · 107 1.8 · 104 ± 2.2 · 104 
Range 4.9 · 106–3.5 · 107 4.5 · 105–3.5 · 107 1.3 · 103–4.8 · 104 

All values are expressed as mean ± standard deviation and range of minimum and maximum. For winter, n = 2; for spring, n = 4 and for summer, n = 4.

Figure 6 shows the efficiencies of pathogen removal and the concentrations of the pathogens in the influent. For the TC, the average reductions were 2.53, 0.87 and 0.44 uLog for winter, spring and summer, respectively. For the FC, average reductions of 2.03, 1.27 and 0.33 uLog were found for winter, spring and summer, respectively. For the somatic coliphages, average reductions of 1.22, 1.18 and 0.58 uLog were found for winter, spring and summer, respectively. For the TC, FC and somatic coliphages, significant differences (p < 0.05) were found between the influent and effluent of the UV system.

Therefore, greater disinfection occurs in winter with greater uLog reduction (>1.22 uLog) for all the microorganisms studied. The elimination of pathogens appears to be directly related to the amount of TSS, and large amounts of TSS are associated with low transmittance values (Figure 5). In this sense, Figure 6 shows that in winter, with an average TSS concentration of 26.7 mg/L, an average reduction of 2.53 uLog in the TC is obtained. In contrast, in summer, an average TSS of 34.7 mg/L is observed, and an average reduction of 0.44 uLog in the TC is obtained.

Figure 6

Efficiencies of elimination of pathogens (bar chart) for (a) total coliforms, (b) fecal coliforms and (c) somatic coliphages and concentration of pathogens in the influent (□) during the studied seasons. For winter, n = 2; for spring, n = 4 and for summer, n = 4.

Figure 6

Efficiencies of elimination of pathogens (bar chart) for (a) total coliforms, (b) fecal coliforms and (c) somatic coliphages and concentration of pathogens in the influent (□) during the studied seasons. For winter, n = 2; for spring, n = 4 and for summer, n = 4.

However, particles larger than 10 μm in size in summer prevented disinfection, protecting the microorganisms due to particle–microorganism association. In the spring, the average particle size is 23.88 nm, with 95% of the particles sized 8.5 μm. Carré et al. (2018) found that the number of particles larger than 25 μm correlates well with the TSS, turbidity and transmittance, affecting disinfection by the dispersion of UV light and protecting the bacteria within their compact nuclei.

Analysis of reactivation of microorganisms

Figure 7 shows the analysis of photoreactivation and reactivation in darkness, which exhibit different trends. For the photoreactivation, a peak is seen at 4 hours, with 27.8% reactivation. This result is higher than that obtained by Zhou et al. (2017) (2.6–9.9%), who measured a peak survival rate at 5 hours under natural light in a sample previously processed by a disinfection system with a potential of 30 mJ/cm2 as in this study.

Increasing the UV dose above 30 mJ/cm2 reduces the concentration of photoreactivated microorganisms as a result of irreversible damage from irradiation (Guo et al. 2013). This effect can be seen in the results obtained by Vélez-Colmenares et al. (2012), where at a dose of 150 mJ/cm2, the survival percentage was less than 0.2%, and when a dose of 57 mJ/cm2 was applied, this value increased to above 1%. Typical photoreactivation behavior due to solar or visible artificial radiation (such as that from lamps) is that after the reactivation phase, a decay phase occurs, similar to that shown in Figure 7. The radiation from this type of lamp, such as the lamp used in this study, simultaneously causes reactivation (UV-A) and inactivation (UV-B and UV-C), as proposed by Šuligoj et al. (2018). This behavior may be associated with two possible consecutive pathways: the first stage is an adequate UV dose that eliminates most of the pathogens and allows only a fraction of the pathogens to reactivate and in the second stage, due to a high accumulation of radiation by pathogens, the pathogens have been inactivated (Giannakis et al. 2015).

Figure 7

Survival ratio of total coliforms exposed to the UV System. Photoreactivation (▪), reactivation in darkness (♦). For photoreactivation n = 2; for reactivation in darkness n = 2.

Figure 7

Survival ratio of total coliforms exposed to the UV System. Photoreactivation (▪), reactivation in darkness (♦). For photoreactivation n = 2; for reactivation in darkness n = 2.

Regarding reactivation in darkness, no peak is seen, similar to the results of Li et al. (2017). Carré et al. (2018) mentioned that spontaneous repair of UV-damaged cells can occur, but this phenomenon is considered insignificant in the dark. These observations show that in the systems studied, even with photoreactivation, the UV system is capable of inducing sufficient damage to practically eliminate pathogens, with only approximately 5% surviving.

CONCLUSIONS

Considering the results of this study, the greatest influences on the disinfection rate (2.03 uLog reduction in the FC) were the low transmittance (<34%), large particle size (>10 μm) and large amount of TSS (>26.7 mg/L) in the system, which favored the ‘shade’ effect.

The TSS concentration directly affected the efficiency of the elimination of microorganisms by the UV system. A reduction of 2.03 uLog in the FC was obtained in winter when the amount of TSS was 26.7 mg/L, while in summer, when the amount of TSS was 34.7 mg/L, a reduction of 0.33 uLog in the FC was obtained.

The photoreactivation analysis showed a peak of 27.8% at 4 hour before moving to a decay phase, decreasing the survival of the microorganisms to 5%. In darkness, no reactivation was observed.

ACKNOWLEDGEMENTS

This work was supported by CONICYT/FONDAP/15130015.

REFERENCES

REFERENCES
APHA/AWWA/WEF
2005
Standard Methods for the Examination of Water and Wastewater
, 21st edn.
American Public Health Association/American Water Works Association/Water Environment Federation
,
Washington, DC
,
USA
.
Azaizeh
H.
Linden
K.
Barstow
C.
Kalbouneh
S.
Tellawi
A.
Albalawneh
A.
Gerchman
Y.
2013
Constructed wetlands combined with UV disinfection systems for removal of enteric pathogens and wastewater contaminants
.
Water Science & Technology
67
(
3
),
651
657
.
doi: 10.2166/wst.2012.615
.
Bertrand-Krajewski
J.
Joannis
C.
Chebbo
G.
Ruban
G.
Métadier
M.
Lacour
C.
2010
Comment utiliser la turbidité pour estimer en continu les concentrations en MES et/ou DCO-Une approche méthodologique pour les réseaux d'assainissement
.
Techniques Sciences Méthodes
105
(
1/2
),
36
46
.
Betancourt
W.
Rose
J.
2004
Drinking water treatment processes for removal of Cryptosporidium and Giardia
.
Veterinary Parasitology
126
(
1–2
),
219
234
.
doi: 10.1016/j.vetpar.2004.09.002
.
Bhattacharjee
S.
2016
DLS and zeta potential–what they are and what they are not?
Journal of Controlled Release
235
,
337
351
.
doi: 10.1016/j.jconrel.2016.06.017
.
Bilotta
G.
Brazier
R.
2008
Understanding the influence of suspended solids on water quality and aquatic biota
.
Water Research
42
(
12
),
2849
2861
.
doi: 10.1016/j.watres.2008.03.018
.
Carré
E.
Pérot
J.
Jauzein
V.
Lopez-Ferber
M.
2018
Impact of suspended particles on UV disinfection of activated-sludge effluent with the aim of reclamation
.
Journal of Water Process Engineering
22
,
87
93
.
doi: 10.1016/j.jwpe.2018.01.016
.
Chahal
C.
Van Den Akker
B.
Young
F.
Franco
C.
Blackbeard
J.
Monis
P.
2016
Chapter two – pathogen and particle associations in wastewater: significance and implications for treatment and disinfection processes
.
Advances in Applied Microbiology
97
,
63
119
.
doi: 10.1016/bs.aambs.2016.08.001
.
Chatterley
C.
Linden
K.
2010
Demonstration and evaluation of germicidal UV-LEDs for point-of-use water disinfection
.
Journal of Water and Health
8
(
3
),
479
486
.
doi: 10.2166/wh.2010.124
.
Das
T.
2001
Ultraviolet disinfection application to a wastewater treatment plant
.
Clean Products and Processes
3
(
2
),
69
80
.
doi: 10.1007/s100980100108
.
Di Rienzo
J.
Casanoves
F.
Balzarini
M.
Gonzales
L.
Tableda
M.
Robledo
C.
2011
InfoStat Statistical Software (Version 8.0)
.
National University of Córdoba
,
Córdoba
,
Argentina
.
Emerick
R.
Loge
F.
Thompson
D.
Darby
J.
1999
Factors influencing ultraviolet disinfection performance part II: association of coliform bacteria with wastewater particles
.
Water Environment Research
71
(
6
),
1178
1187
.
doi: 10.2175/106143097X122004
.
Giannakis
S.
Darakas
E.
Escalas-Cañellas
A.
Pulgarin
C.
2015
Environmental considerations on solar disinfection of wastewater and the subsequent bacterial (re)growth
.
Photochemical & Photobiological Sciences
14
(
3
),
618
625
.
doi: 10.1039/C4PP00266 K
.
Guo
M.
Huang
J.
Hu
H.
Liu
W.
Yang
J.
2013
Quantitative characterization and prediction modeling of photoreactivation of coliforms after ultraviolet disinfection of reclaimed municipal wastewater
.
Water, Air, & Soil Pollution
224
(
11
),
1774
.
doi: 10.1007/s11270-013-1774-z
.
Hannouche
A.
Chebbo
G.
Ruban
G.
Tassin
B.
Lemaire
B.
Joannis
C.
2011
Relationship between turbidity and total suspended solids concentration within a combined sewer system
.
Water Science and Technology
64
(
12
),
2445
2452
.
doi: 10.2166/wst.2011.779
.
Hassard
F.
Gwyther
C.
Farkas
K.
Andrews
A.
Jones
V.
Cox
B.
Brett
H.
Jones
D.
McDonald
J.
Malham
S.
2016
Abundance and distribution of enteric bacteria and viruses in coastal and estuarine sediments – a review
.
Frontiers in Microbiology
7
,
1692
.
doi: 10.3389/fmicb.2016.01692
.
Jolis
D.
Lam
C.
Pitt
P.
2001
Particle effects on ultraviolet disinfection of coliform bacteria in recycled water
.
Water Environment Research
73
(
2
),
233
236
.
doi: 10.2175/ 106143001X139218
.
Leiva
A.
Albarrán
A.
López
D.
Vidal
G.
2019
Evaluation of phytotoxicity of effluents from activated sludge and constructed wetland system for wastewater reuse
.
Water Science & Technology
79
(
4
),
656
667
.
doi: 10.2166/wst.2019.093
.
Li
G.
Wang
W.
Huo
Z.
Lu
Y.
Hu
H.
2017
Comparison of UV-LED and low pressure UV for water disinfection: photoreactivation and dark repair of Escherichia coli
.
Water Research
126
,
134
143
.
doi: 10.1016/j.watres.2017.09.030
.
López
D.
Fuenzalida
D.
Vera
I.
Rojas
K.
Vidal
G.
2015
Relationship between the removal of organic matter and the production of methane in subsurface flow constructed wetlands designed for wastewater treatment
.
Ecological Engineering
83
,
296
304
.
doi: 10.1016/j.ecoleng.2015.06.037
.
López
D.
Leiva
A. M.
Arismendi
W.
Vidal
G.
2019
Influence of design and operational parameters on the pathogens reduction in constructed wetland under the climate change scenario
.
Reviews in Environmental Science and Bio/Technology
18
(
1
),
101
125
.
doi: 10.1007/s11157-019-09493-1
.
Masbough
A.
Frankowski
K.
Hall
K.
Duff
S.
2005
The effectiveness of constructed wetland for treatment of woodwaste leachate
.
Ecological Engineering
25
(
5
),
552
566
.
doi: 10.1016/j.ecoleng.2005.07.006
.
Pavanelli
D.
Pagliarani
A.
2002
SW – soil and water: monitoring water flow, turbidity and suspended sediment load, from an Apennine catchment basin, Italy
.
Biosystems Engineering
83
(
4
),
463
468
.
doi: 10.1006/bioe.2002.0126
.
Šuligoj
A.
Arčon
I.
Mazaj
M.
Dražić
G.
Arčon
D.
Cool
P.
Tušar
N.
2018
Surface modified titanium dioxide using transition metals: nickel as a winning transition metal for solar light photocatalysis
.
Journal of Materials Chemistry
6
(
21
),
9882
9892
.
doi: 10.1039/c7ta07176 k
.
Vázquez-López
M.
Amabilis-Sosa
L. E.
Moeller-Chávez
G. E.
Roé-Sosa
A.
Neumann
P.
Vidal
G.
2019
Evaluation of the ultrasound effect on treated municipal wastewater
.
Environmental Technology
40
(
27
),
3568
3577
.
doi: 10.1080/09593330.2018.1481889
.
Vélez-Colmenares
J.
Acevedo
A.
Salcedo
I.
Nebot
E.
2012
New kinetic model for predicting the photoreactivation of bacteria with sunlight
.
Journal of Photochemistry and Photobiology B: Biology
117
,
278
285
.
doi: 10.1016/j.jphotobiol.2012.09.005
.
Vera
I.
García
J.
Sáez
K.
Moragas
L.
Vidal
G.
2011
Performance evaluation of eight years experience of constructed wetland systems in Catalonia as alternative treatment for small communities
.
Ecological Engineering
37
(
2
),
364
371
.
doi: 10.1016/j.ecoleng.2010.11.031
.
Vidal
G.
Nieto
J.
Mansilla
H. D.
Bornhardt
C.
2004
Combined oxidative and biological treatment of separated streams of tannery wastewater
.
Water Science and Technology
49
(
4
),
287
292
.
doi: 0000-0001-7433-5004
.
Vymazal
J.
2007
Removal of nutrients in various types of constructed wetlands
.
Science of the Total Environment
380
(
1–3
),
48
65
.
doi: 10.1016/j.scitotenv.2006.09.014
.
Yeber
M. C.
Soto
C.
Riveros
R.
Navarrete
J.
Vidal
G.
2009
Optimization by factorial design of copper (II) and toxicity removal using a photocatalytic process with TiO2 as semiconductor
.
Chemical Engineering Journal
152
,
14
19
.
doi: 10.1016/j.cej.2009.03.021
.
Zhou
H.
Smith
D.
2002
Advanced technologies in water and wastewater treatment
.
Journal of Environmental Engineering and Science
1
(
4
),
247
264
.
doi: 10.1139/s02-020
.
Zhou
X.
Li
Z.
Lan
J.
Yan
Y.
Zhu
N.
2017
Kinetics of inactivation and photoreactivation of Escherichia coli using ultrasound-enhanced UV-C light-emitting diodes disinfection
.
Ultrasonics Sonochemistry
35
(
Part A
),
471
477
.
doi: 10.1016/j.ultsonch.2016.10.028
.