## 5 Discussion  
The groundwater radon concentrations in sites 1 and 2 were used to estimate average values for all monitoring wells over the study period. Radon data from clean wells of site 1 (those with higher radon levels and nearly no NAPLs, PZ03, PZ07, PZ14, PZ25, and PZ26) were averaged and used as a reference for calculating the radon deficit of dirty wells, as the ratio of the radon concentration in NAPL polluted aquifer and the equilibrium radon concentration in clean groundwater. The latter (119 \(\pm\) 11 Bq L\({}^{-1}\)) was compared and verified with a corresponding value (\({}^{222}\)Rn\({}_{\rm C}\), 129 \(\pm\) 10 Bq L\({}^{-1}\)) calculated from the average \({}^{226}\)Ra specific activity of the Alban Hills Hydrogeological Unit (A\({}_{\rm Ra}\), 85 Bq kg\({}^{-1}\), see Table 3), according to Equation (2), using a mineral density (r,) of 2.65 kg/L, a porosity (n) of 0.35, and an emanation radon efficiency (E) of 0.2 [23]:  
\[{}^{222}\rm{Rn_{C}}=A_{\rm Ra}\times r\times E/n. \tag{2}\]  
Since the two values agree within the error range, it was possible to determine with good approximation the NAPL saturation of contaminated wells in site 1 according to Equation (1) and show areas with higher saturation in a schematic map (Figure 5a). The well PZ13 and the adjacent area where the residual source zone is presumably located were characterized by the highest NAPL saturation (2.39%), while the wells located down-gradient (PZ02, PZ04, PZ09, and PZ19) showed a saturation around or above 1%, apart from PZ19. PZ19 well was included in the plume zone, even if its saturation was much lower (0.42), because its apparent low saturation is calculated from a radon concentration affected by the proximity of PZ19 to a recharge well. Actually, the introduction of water treated with activated carbon filters absorb, not only the NAPLs, but also radon, which may not have time to reach the full equilibrium with its parent \({}^{226}\)Ra in the rock aquifer (low residence time). Other wells surrounding the main plume or working as pumping wells were less polluted (PZ06, PZ8, PZ17, PZ18, and PZ19). The plume was elongated in the direction of the groundwater flow (NNW-ESE) and was perfectly retained in the fueling station area. Since the monitoring of the water table elevation and the groundwater quality was not continuous but periodical, it was difficult to demonstrate direct relationships between the rainfall distribution and the depth of the groundwater table on one side and the dissolved NAPL on the other. Using factorial analysis can help clarify these interactions. Moreover, a recent study [22] carried out in a nearby area with a similar environmental and hydrogeological setting invoked a possible mechanism of contaminant (MTBE) remobilization from soil grains (mainly zeolites) linked to rainfall washing the terrain. Using the approach presented in [33], a dissipation half-life of 23 days was estimated for MTBE in the groundwater [22]. A similar process could be applied to site 1 where residual NAPLs, located at a depth of 6-12 m, could be removed, dissolved, transported, and affected by degradation and volatilization phenomena (natural attenuation). To assess and quantify these processes, principal component analysis was applied to the groundwater data of site 1, but not to the soil variables that were not available with the same frequency. The testing provided three main factors accounting for 85% of the total variance (Table 4). It is worth noting that the interpretation of factors is not always based on very high scores, but it is still supported by other descriptors. Factor 1 explaining the 50% of variance could be interpreted as the radon deficit process, which is the preferential solubility of radon in the NAPLs. A negative score of the groundwater depth (corresponding to a rainy period) was calculated with a negative value of radon and positive scores for MTBE, ETBE, and the total hydrocarbon concentrations.