## 1 Introduction  
Contamination of groundwater due to environmentally unsustainable human activities has become one the most important pollution threats worldwide. The main sources of pollutants are derived from industrial and agricultural activities. Thus, many inorganic and organic compounds, including pesticides, have increased their presence in surface and groundwater bodies. Pollutants have the potential to modify groundwater chemical composition as well as groundwater ecosystems [1; 2]. Nitrogen-based compounds, such as fertilizers or pesticides, have widespread use in agricultural regions. They may promote eutrophication of surface water bodies and, due to their toxicity, represent a risk for both human health and the survival of aquatic species [3]. Atrazine is within this group of compounds. It is one of the most widely applied herbicides all over the world since its introduction in the 1950s. The persistence of atrazine, its moderate water solubility, and high movement through soil and water key to its potential role as a pollutant in water bodies [4; 5; 6]. Atrazine concentrations above the allowed limits of 0.1 \(\mu\)g/L, established by the European Union [7], have been found in drinking water [8]. In addition, its disruptor endocrine activity is demonstrated with pernicious potential effects on human [9; 10] and animal health [11; 12]. As a consequence, atrazine application was prohibited by the European Union in 2004, although in Spain it was extended until 2007. Even after more than 15 years of being banned in EU and Spain, nowadays considerable concentrations of atrazine and its metabolites are still detected in surface and groundwater bodies due to their elevated persistence in the environment [13]. Over the past few decades, researchers' concerns about atrazine removal paths have been raised regarding its environmentally persistence and water quality impact [14; 15; 16; 17; 18]. Biological processes are the predominant atrazine degradation pathways. [PERSON] et al. [19] revised and explained the enzymes and genes involved in microbial s-triazine metabolism. The reactions occur in an obligate order with funneling of broad-specificity enzymes, encoded by _atzA,B,C/trzN_ genes, into cyanuric acid as an intermediate product. Then, subsequently narrow-specificity enzymes, codified by _atzD,E,F/trzD,F_ genes, finish the s-triazine ring cleavage. The AtzA/TrzN enzymes start a hydrolytic dechlorinating reaction to produce the hydroxyapatite molecule, followed by a sequential and stepwise alkylamino chains removal carried out by AtzB and AtzC enzymes to produce N-isopropylammelide and cyanuric acid, respectively. Then, AtzD/TrzD enzymes hydrolyze cyanuric acid to a biuret intermediate, which is hydrolyzed to allophanic acid by the AtzE enzyme. Finally, allophanic acid is the single hydrolyzing substrate for the AtzF enzyme to produce carbon oxide and ammonia. Microbial communities within aquifers are expected to consist largely of heterotrophs and chemoautotrophs well adapted to the nutrient-poor and oligotrophic groundwater environment [20]. Moreover, under an environment with a long history of repeated xenobiotic exposition, microorganisms are expected to adapt their metabolic mechanisms progressively to carry degradative genes for foreign pollutants [21; 22; 23]. In many deep aquifers, which contain huge water volumes with slow turnover times, in situ bioremediation performed by the local communities of microorganisms is very interesting and sometimes, the only decontamination recourse [24]. This study was carried out to evaluate the presence of bacteria related to atrazine degradation in deep zones of the Petrola hypersaline lake-aquifer system (southeastern Spain). The first molecular assessment of the groundwater bacterial community structure has been made by the sequencing of 16S rRNA gene clones obtained from groundwater samples collected at 37.9 m deep.