Analysis of the Genetic Effects to Frogs (Fejervarya limnocharis) After Acute Lead Exposure In Vivo

This study aimed to investigate the bioaccumulation and elimination of lead (Pb) in Fejervarya limnocharis frogs as well as to determine the genotoxic effects of direct Pb exposure at different concentrations and lengths of time. Four varying concentrations (0, 5, 10 and 20 mg.L-1) of lead acetate (Pb(CH3COO)2) solutions were injected intraperitoneally into F. limnocharis. The concentration of Pb in the water samples used to house the frogs and the concentration of lead in frog muscle tissues were analysed at 24, 48 and 72 hours after injection by inductively coupled plasma optical emission spectrometry. Pb was detected at a level that exceeded the standard (0.03 mg.kg-1) in all samples of frogs injected with Pb. The water samples indicated that the Pb concentrations were significantly different from the control (p < 0.05), except for the 5 and 10 mg.L-1 concentrations after 24 hours groups and 5 mg.L-1 concentration after 48 hours group. Only the concentration of the water in the 20 mg.L-1 for 72 hours group exceeded the standard (0.05 mg.L-1). Genetic differentiation was studied by inter simple sequence repeats (ISSR) with dendrogram construction and analysis of genetic similarity (S) for each duration of exposure. A total of 1158, 1205 and 1277 bands were generated by ISSR for the 24, 48 and 72 hours groups, respectively. In each dendrogram, individual injections with the same Pb concentration clustered together, and it appeared that higher concentrations resulted in greater genotoxicity. Genotoxicity was concentrationand time-dependent, with a correlation between the concentration and S-value for the 72 hours group (R2 = 0.77, p < 0.05). In addition, this study could provide a basic application to develop F. limnocharis as a biomarker for Pb contamination by measuring genotoxic consequences.


INTRODUCTION
Lead (Pb) contamination is an important environmental problem that affects water, soil and air quality worldwide (Mitchell et al. 2011, Jaishankar & Mathew 2014. Pb contamination is especially problematic near mines, at sites of industrial wastewater discharge and from leachate from polluted areas. Pb is detrimental to various animals, and its toxicity can harm human health as the accumulation of Pb reaches excessive levels (IARC 2004, Martin & Griswold 2009, Olafisoye et al. 2013. Pb can enter creatures through many routes, including skin contact, consumption of contaminated foods and water and breathing contaminated air (Kakker & Jaffery 2004, Yan et al. 2018). Low-level Pb toxicity is often nonspecific and may cause fatigue and intellectual impairment (Méndez-Gómez et al. 2008). Severe Pb toxicity affects multiple organ systems and can result in renal, haematological, neurological and cardiovascular diseases (Correia et al. 2000, Kalia & Flora 2005, Navas et al. 2007, ATSDR 2007. Pb toxicity can occur in kidney, spleen, heart, liver, lung, bone, brain and skeletal muscle (Patra et al. 2001, Vargas et al. 2003, ATSDR 2007, Sharma et al. 2011. Acute and chronic Pb toxicity in organisms depends on the concentration and duration of exposure (Needleman 2004, ATSDR 2007, Cleveland et al. 2008).
Furthermore, several published reports have revealed that Pb causes genotoxicity and cytotoxicity through oxidative stress and DNA damage (Kasuba et al. 2004, Zhang & Li 2013. The genotoxicity of Pb is related to the oxidative stress that occurs when there is a lack of proportion between the formation and removal of reactive oxygen species (ROS). During oxidative stress, ROS impair cells by oxidizing membrane lipids and proteins as well as DNA. DNA damage is a process with a facilitative role for Pb in carcinogenesis and cell death (Kosnett 2006, Beyersmann & Hartwig 2008.
Polymerase chain reaction with inter simple sequence repeats (PCR-ISSR) is a technique that can be used to analyze genetic relationships using molecular markers and has high resolving power at the population level without prior molecular knowledge of the organism (Zietkiewicz et al. 1994, Wu et al. 1994. This technique can be used to detect various types of genetic differentiation or DNA damage and mutations, such as point mutations, rearrangements and small deletions or insertions of DNA in plants and animals (Neeratanaphan et al. 2014, Mahfouz & Rayan 2016, Al-Qurainy et al. 2017. The ISSR method represents one of the most promising tools for the detection of genetic alterations in response to heavy metal toxicity by looking directly at the DNA sequence and structure (Behura 2006, Héry et al. 2008. Amphibians are sensitive and important indicators to exposure to environmental pollutants because of their extremely permeable skin (Huang et al. 2007, Falfushinska et al. 2008, Kerby et al. 2010, Burlibasa et al. 2011. The rice field frog (Fejervarya limnocharis) is a native species found in Southeast Asia, including Thailand, Laos and Cambodia (Liu et al. 2011), and this creature is an important animal in the food chain, particularly on agricultural land in Southeast Asia. The toxicity of Pb has been studied widely in fish, but evidence relating to amphibians is relatively sparse (Linder & Grillitsch 2000). Some studies have related the concentration of Pb in amphibians to that in their surrounding environment; however, these field studies cannot establish causes and effects. Therefore, laboratory-based studies are required to reveal the effects of Pb accumulation in F. limnocharis by simulation of environmental exposure. A frog (F. limnocharis) model was chosen for this study because this frog species is abundant in aquatic ecosystems. The objectives of the study were to investigate the concentrations of Pb in F. limnocharis and water after in vivo injection with different concentrations of lead acetate for different exposure times and to determine the relationship to genomic changes using the PCR-ISSR technique.

MATERIALS AND METHODS
Experimental design: Frogs of the species Fejervarya limnocharis, weighing approximately 15 grams, were collected from a private culture farm located in Khon Kaen Province, Thailand. Healthy F. limnocharis were randomly selected for the experiment. The frogs were allowed to acclimate to the laboratory conditions for 5 days and then randomly divided into 12 experimental groups: 3 control groups and 9 treatment groups. The experiments were performed in triplicate with 5 frogs in each group. Each plastic bucket as the experimental unit had a 15 L capacity and contained 1 L of dechlorinated tap water.
Lead acetate exposure concentration: For the lead acetate exposure study, 100 µL of deionized water (control groups 1-3) or solutions of Pb(CH 3 COO) 2 (experimental groups 4-12) of varying concentrations (5, 10 and 20 mg.L -1 ) were prepared and applied to the frogs. At each concentration, the frogs were exposed to the injected solution for 24, 48 or 72 hours before being analysed. The water in the buckets was not changed during the experimental period.
Lead acetate analysis in frogs (F. limnocharis): Frogs of the species F. limnocharis were anaesthetized with ice, and the muscle samples were cut into small pieces. Exactly 1 g of muscle tissue was placed into a beaker and digested with 7 mL of concentrated nitric acid (HNO 3 ) and 1 mL of concentrated hydrogen peroxide (H 2 O 2 ), placed in a water bath at 90 ± 5°C for 2 hours, and then allowed to cool to room temperature. The digested muscle samples were adjusted to a final volume of 25 mL with deionized water and then filtered.

Inductively coupled plasma optical emission spectrometry (ICP-OES) analysis:
The digested samples of water and frog muscle were analysed for their Pb concentrations using ICP-OES. The ICP-OES wavelength for Pb analysis was set at 220.353 nm. The detection limit of Pb was 0.005 mg.L -1 . The precision of the metal concentrations was evaluated with a certified reference material (CRM) via the 3111C method (APHA 2005, Chand & Prasad 2013).
Genotoxicity analysis: Total genomic DNA was extracted from F. limnocharis livers using a Genomic DNA Extraction Kit (Tissue) (RBC Bioscience) following the kit protocol. Extracted DNA was inspected by 0.8% agarose gel electrophoresis and ViSafe Green gel staining.
Later, a polymerase chain reaction was performed with a reaction mixture (10 µL) consisting of BlueMix DNA polymerase MasterMix (RBC Bioscience), 1.25 µM primers and 20 ng of genomic DNA extracted from F. limnocharis liver tissue. Twenty-four ISSR primers were screened with DNA from the representative frogs at the different concentrations and at each duration length, and the experimental primers were selected based on their amplification profile. The 14 primers that amplified favourably for the 24, 48 and 72 hours experiments are shown in Table 1.
The PCR profile was demonstrated by predenaturation at 94°C for 5 minutes, 35 cycles of denaturation at 94°C for 1 minute, annealing at 52°C for 1 minute and 30 seconds, and extension at 72°C for 2 minutes, and a final extension at 72°C for 7 minutes using a thermal cycler (Analytik Jena). After amplification, the PCR products were resolved by electrophoresis on 2% agarose gels and ViSafe Green gel staining.

Data analyses:
The Pb concentrations in the water and frog samples were analysed using ANOVA followed by Tukey's post hoc test. The results are expressed as the mean ± SD. All statistical data were carried out with SPSS for Windows (version 19), and the significance level was set at p < 0.05.
Polymorphic bands amplified by the ISSR markers were scored as present =1 or absent = 0 compared to a standard DNA ladder. The data were entered into NTSYS-pc version 2.10p (Rohlf 2009) to construct the dendrogram and perform statistical analysis. Regression analysis was performed using SPSS.

Pb concentrations in the water and F. limnocharis samples:
The Pb concentrations in the water after F. limnocharis were injected with Pb concentrations of 0, 5, 10 and 20 mg.L -1 for exposure times of 24, 48 and 72 hours are given in Table 2. The Pb concentrations of the water samples did not exceed the standard of the Pollution Control Department of Thailand of water quality standards (0.05 mg.L -1 ) (TPCD 1986), except for the highest Pb concentration of the 20 mg.L -1 for 72 hours (0.05 ± 0.01 mg.L -1 ). The Pb concentrations of the water samples between the experimental and control groups were significantly different (p < 0.05) for all experimental groups except for the 5 and 10 mg.L -1 groups for 24 hours and 5 mg.L -1 group for 48 hours.
The Pb concentrations in F. limnocharis muscle after injection with different Pb concentrations and durations of exposure are given in Table 3. Pb was not detected in the F. limnocharis muscles of the control group. The Pb concentrations of F. limnocharis muscles in all experimental groups exceeded the standard of the Pollution Control Department of Thailand (0.03 mg.kg -1 ) (TPCD 2001). The highest Pb concentration was found in the 20 mg. L -1 72 hours group (0.78 ± 0.04 mg.kg -1 ). The statistical analysis results indicated that the Pb concentrations in all experimental F. limnocharis muscle samples were significantly different from the controls (p < 0.05).
These results suggest that the Pb found in the water samples originated from F. limnocharis frogs. It appeared that higher concentrations of injected Pb and longer exposure times resulted in increased Pb concentrations in the water. Pb accumulates mainly in the liver and kidney of F. limnocharis (Hopkins 1989, Jančová et al. 2002, Zhang et al. 2007) and is excreted through the urine, faeces and skin (Akerstrom et al. 2013). Consequently, F. limnocharis may not eliminate all of the Pb from their bodies, and some Pb remains in the muscles and other organs.
Several studies are confirming that the Pb concentrations absorbed in frog bodies can be transported to various target organs, such as the liver and kidney, compared with transportation to the muscles (Jarup et al. 1988, Grosicki & Kowalski 2002, Formicki & Stawarz 2004. The intracellular movement of Pb across cellular membranes occurs via active transport, which is facilitated by P-type ATPase ionic pumps (Zelikoff et al. 1988, Gatti et al. 2000, Patric 2006). Among heavy metals, Pb has a higher propensity to bioaccumulate in frog organs involved in the detoxification process.
Genotoxic effects in F. limnocharis: Fourteen clear and reproducible ISSR primers generated ISSR bands for the 24, 48 and 72 hours groups as 1158, 1205 and 1277, respectively, ranging in size from 100 bp to 3000 bp. Examples of these ISSR bands for different primers are shown in Fig. 1. These bands were used for dendrogram construction for each duration of exposure (Figs. 2-4). For each dendrogram, the closest individuals from the same Pb concentration clusters demonstrate that they are genetically similar. These clusters were grouped at varying distances between the other Pb concentration groups with a trend towards greater distances as greater Pb concentrations.  14 ISSR primers amplified differentiation values of the experimental groups compared with the controls were 0.55, 0.44-0.45 and 0.48 for 5, 10 and 20 mg.L -1 , respectively. At 72 hours of exposure, the differentiation values of the samples from each experimental group compared to the control group were 0.69, 0.59-0.60 and 0.49 for 5, 10 and 20 mg.L -1 , respectively. As Pb was detected in all F. limnocharis samples except for the control group, this suggests that the injected Pb was responsible Among heavy metals, Pb has a higher propensity to bioaccumulate in frog organs involved in the detoxification process.   (Hopkins 1989, Jančová et al. 2002, Zhang et al. 2007) and is excreted through the urine, faeces and skin (Akerstrom et al. 2013). Consequently, F.
limnocharis may not eliminate all of the Pb from their bodies, and some Pb remains in the muscles and other organs.
Several studies are confirming that the Pb concentrations absorbed in frog bodies can be transported to various target organs, such as the liver and kidney, compared with transportation to the muscles (Jarup et al. 1988, Grosicki & Kowalski 2002, Formicki & Stawarz 2004. The intracellular movement of Pb across cellular membranes occurs via active transport, which is facilitated by P-type ATPase ionic pumps (Zelikoff et al. 1988, Gatti et al. 2000, Patric 2006).
Among heavy metals, Pb has a higher propensity to bioaccumulate in frog organs involved in the detoxification process.

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as biomarkers of the genotoxic effects of the environment is that in vivo intraperitoneal injections do not completely simulate exposure to Pb in the natural environment. Therefore, this study provides basic knowledge for future research, particularly for the use of frogs as biomarkers of heavy metal contamination in relation to genotoxicity consequences. concentration differences and genetic differentiation values for 24 hours (R 2 = 0.05, p = 0.89) (Fig. 5) and 48 hours (R 2 = 0.086, p = 0.57) of exposure (Fig. 6). For 72 hours of exposure, there was a significant correlation between the differences in Pb concentration and genetic similarity values (R 2 = 0.77, p < 0.05) (Fig. 7). These results suggest that the for the genotoxic effects, as evidenced by the genetic differentiation values of the experimental groups compared with the control group. Differences in Pb concentrations were calculated between the groups and correlated against the genetic differentiation values with linear regression analysis. There was no significant correlation between the            genotoxic effects of Pb depend on both the exposure time and concentration.
The results of this study indicate that intraperitoneal injection of Pb solutions leads to contamination of the water in the frog's environment by excretion via urine and faeces. This study also showed that Pb induced genotoxicity in a time and concentration-dependent manner. These results have also been observed in many other studies; however, the vast majority of previous studies relating to the genotoxic effects of Pb have been undertaken in fish, rats, mice and rabbits, with a clear gap in the literature regarding frogs (García-Lestón et al. 2010). There are various mechanisms of genotoxicity that have been discussed in previous literature, including oxidative stress from the generation of reactive oxygen species and by DNA repair enzymes (Hwang & Kim 2007, Gueranger et al. 2011, Krisko & Radman 2013, McAdam et al. 2016). The reported mechanisms may be responsible for the genotoxic effects of Pb observed in the frogs in this study.
Frogs were selected for this study because they have the potential to be useful biomarkers for environmental contamination of Pb due to their highly permeable skin (Othman 2009;Thammachot et al. 2012). The disadvantage of this study in exploring the application of frogs as biomarkers of the genotoxic effects of the environment is that in vivo intraperitoneal injections do not completely simulate exposure to Pb in the natural environment. Therefore, this study provides basic knowledge for future research, particularly for the use of frogs as biomarkers of heavy metal contamination in relation to genotoxicity consequences.

CONCLUSION
Intraperitoneal injection of a Pb solution into F. limnocharis frogs was responsible for the contamination of lead into water and its accumulation in muscles. Genotoxicity effects after Pb exposure, as evidenced by dendrograms and genetic differentiation values, are concentration-dependent and additive. This study revealed that F. limnocharis frogs can potentially be used as biomarkers for Pb contamination through measurement of the genotoxic effects.

ACKNOWLEDGMENTS
This research was supported by the Research Group on Toxic Substances in Livestock and Aquatic Animals and the Gradnuate School of Khon Kaen University, Thailand.

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c differentiation values according to differences in the concentration for 24 hours of exposure.
c differentiation values according to differences in the concentration for 48 hours of exposure.