3.1. General Evaluation of Parameters Measured in Groundwater and Leachate
The content of HMs in groundwater and leachate can be evaluated from several perspectives both as the mean values for individual years (
Figure 3) and as the mean values for the whole period of monitoring (2008–2013) (
Figure 4A,B). The measured values revealed several results. Firstly, more profound concentrations of HMs (
Figure 3A) were found only in Ni, Zn and Cr, with the values regularly exceeding the limit of 0.1 mg L
−1, which could be considered significant with respect to the sensitivity of methods used. Other HMs such as, Pb, Hg and Cd (
Figure 3B) were most of the time present in very low concentrations < 0.07 mg L
−1. Secondly, no significant differences (P < 0.05) were observed among the respective monitoring drill holes (
Figure 4), only between all drill holes and the LP. Thirdly, neither the date nor the year of sampling had a significant influence on the concentrations of the selected HMs in the samples. This was also corroborated by the ANOVA results (
Appendix A Table A1 and
Table A2) which did not indicate any significant differences. The measured values showed only partial differences between the individual drill holes and the well. By contrast, the concentration of HMs measured in the LP reached demonstrably (
P < 0.05) the highest values.
Increased concentrations were found only in three HMs: Ni, Zn and Cr. The other HMs were on the limit of determination. Based on the previous scientific studies of [
27,
28,
29], the concentration of HMs was chosen as a major indicator of possible adverse effects of the hypothetical leakage of contaminated leachates from the landfill body. The values of Ni, Zn and Cr measured in the monitoring drill holes (MV-1, MV-2A, MV-4, MV-5 and the well) indicated that groundwater contamination with HMs was not increased. Only one significant difference was found in the concentration of Ni between MV-4 and the remaining drill holes. The measured values were in accordance with those found by Olafisoye et al. and Chen et al. [
30,
31], who consider Ni, Zn and Cr concentrations <0.1 mg L
-1 acceptable and not representing a profound danger to the surrounding environment. Furthermore, it can be assumed that the number of leachates will decrease with the shrinking area of the active landfill part (
Figure 2), which was reduced by more than 5% in the period from 2000 to 2014. The reason for this is that the reclaimed part of the formerly active landfill is able to retain a greater amount of precipitation in the landfill surface layers (because of the vegetation) and the amount of leachates will be lower than in the open landfill body.
The below comparison with the World Health Organization (WHO) limits (2011 and 2017) [
32,
33] in
Table 3 shows clearly that the measured Ni, Zn and Cr concentrations were higher in the LP when compared with those in the drill holes and in the shaft ring well. These increased HMs concentrations in the measuring point of LP demonstrate that the mechanisms preventing HMs leaks from the landfill into the surrounding environment function properly. The most important measure is seen in the landfill insulation layer, which prevents HMs leakage into groundwater. The LP water exhibited the highest concentrations of Ni, Zn and Cr (ranging from 0.3 to 0.9 mg L
-1) in the respective years of monitoring (
Figure 3; 2008–2013) as well as average concentrations of HMs for the whole monitoring period (
Figure 4; 2008–2013).
The contents of P and N (
Figure 5) and COD, AOX and PAH were also monitored (
Figure 6).
These supplementary indicators may point to other potential adverse effects on the environment in the case of possible leakages. Based on the presence of AOX and PAH in the leachates, the origin of the HMs can be determined because, for example, AOX occur most frequently in leachates that have come into contact with cellulose- and ash-based waste from MSW incinerators [
34,
35]. On the other hand, the values of COD and the contents of biogenic elements (N and P) may indicate the organic origin of leachate pollution. The values of P
tot content ranged up to 10 mg L
−1 with no significant differences among the respective monitoring points. Based on these values, it can be stated that runoff from adjacent fields did not reach the landfill body (
Figure 1;
Appendix A Figure A1), which is also indicated by runoff lines directed away from the landfill. Another reason may be the generally low mobility of P substances [
36,
37,
38] inhibiting their possible washout from the landfill body in the case of their presence therein. In the drill holes and shaft ring well, the mean N
tot value did not exceed the boundary of 60 mg L
−1; by contrast, in the MV-2A, MV-4, MV-5 monitoring points and in the well, it dropped below 50 mg L
−1. For comparison, the Drinking Water Directive of the European Union (EU) (98/83/EC) sets the value of 50 mg L
−1 as the limit for the content of nitrates in drinking water. The value of N
tot in the above-mentioned sampling points did not exceed the set limit with the exception of MV-1, where the detected concentration was 58 mg L
−1. These values indicate that there were no leakages of organic pollutants from the landfill body. Moreover,
Appendix A Figure A2 contains information about the concentration of the individual form of N and nitrate N, which represents more than 95% of the N
tot within all variants except for the LP. On the contrary, the analysis of samples for the period from 2008 to 2013 revealed an increased content of N
tot in the LP. And only 10% of this content was formed by mineral forms of N. This indicates that the LP held not only the leachate from the landfill but also surface flushes from the individual landfill parts. Together with the concentrations of HMs, this value confirmed once again that the landfill had been designed properly, and possible leachates were captured by the insulation layer and drained into the LP. This was also corroborated by the COD values, which were minimal in all monitoring objects (< 3 mg L
−1) except for the LP, where they ranged from 140 to 490 mg L
-1 in the studied period. The COD parameter serves to determine an ideal amount of O
2 that would be required for oxidation of all organic substances in water. Thus, its value is directly proportional to the content of potential OP indicators (Oxygen indicators: COD; AOX and Polyromantic indicator: PAH) in water [
39,
40,
41]. In the EU, the limits for drinking water and treated wastewater are 3 mg L
−1 and 55 mg L
−1, respectively (Directives 91/271/EEC and 2000/60/EC). Natural sources mostly exhibit values from 100 to 60,000 mg L
−1 [
39,
40,
41]. Therefore, we can claim that the COD values measured in the monitoring points exhibited no contamination by pollutants of organic origin.
The AOX and PAH contents were monitored, too (
Figure 6). The AOX parameter indicates the content of organic substances binding some elements from the group of halogens. These substances are potentially toxic, especially for aquatic ecosystems [
35,
40]. A description of their impact on human health is complicated because AOX represents a wide group of substances with diverse impacts [
41]. The average values measured in all drill holes and in the well did not exceed 60 µg L
−1. The values measured in the respective monitoring objects (with the exception of LP) ranged from a minimum of 8.5 to a maximum of 232 µg L
−1 with no statistically significant differences among the monitoring objects. There was no evidence that the values would be directly affected by waste landfilling. The substances of the AOX group can be generated naturally, too, for instance by enzymes produced by microorganisms [
42].
As to the AOX concentration, an exception was the LP, in which the mean value for the period from 2008–2013 reached 1000 µg L
−1 and the values detected during the monitoring ranged from 650 to 1260. The PAH content was also monitored. An increased PAH concentration would have indicated the increased presence of waste from powerplants and other incineration facilities and asphalt mixtures in the landfill body [
43]. This was not confirmed because the measured PAH values were at the limit of detection, i.e., around 0.09 mg L
−1. No differences were observed among the monitored objects.
The measured concentrations of HMs exceeded the value of 0.01 mg L
−1 only in Ni, Zn and Cr. The increased values (> 1 mg L
−1) were detected only in the LP. The monitoring drill holes as well as the shaft ring well exhibited at all times the values of these HMs below 0.2 mg L
−1, which represented no risk [
27]. It is important to take into account the landfill location on an elevated site in the given locality (
Figure 1;
Appendix A Figure A1). Should the landfill construction have some defects, namely in terms of insulation (the prevention of leachate leakage), the transportation of measured substances (HMs, PAH, AOX) into the surrounding lands would have occurred [
44] because all runoff lines are directed into the adjacent farmlands (
Appendix A Figure A1). The direction of surface and sub-surface runoffs is plotted in
Figure 1 and in
Appendix A Figure A1 as blue dot-dash arrows. Nevertheless, the measured values disproved such a scenario because the monitoring drill holes did not exhibit increased values of the parameters.
3.2. Relationship between the Individual Measured Parameters
The strength of the potential relations among the respective parameters was determined through the correlation analysis at a significance level of
P < 0.05. The correlation matrix of data obtained from the PCA is shown in
Appendix A Table A3. The PCA analysis results are given in
Appendix A Table A4. A graphical representation of the correlation of each parameter is given in
Appendix A Figure A3. There was a significant and positive relationship between pH and P
tot, N
tot, Cd, AOX, Zn, Cr, Pb, Ni, COD, As, PAH (r = 0.505–0.877). There was also a significant and positive correlation between P
total and N
total, Zn, Cr, Pb, Ni, As, COD, PAH, AOX (r = 0.501–0.968). A clear dependence was observed of the presence of organic substances in the leachate and the presence of most HMs. In particular, a significant and positive relationship was found between AOX and pH, P
tot, N
tot, Cd, Zn, Cr, Pb, Ni, As, COD, PAH, (r = 0.508–0.963). These significant correlations again indicate a relationship between the organic sources of pollution and the other pollutants (AOX and HMs). Another significant and positive relationship was detected between Zn and pH, P
tot, N
tot, AOX, Zn, Cr, Pb, Ni, As, COD, PAH, (r = 0.578–0.834). Furthermore, all measured values confirmed that the AOX concentration correlated with leachate characteristics like pH, P
tot and N
tot content. Also, a strong relationship was found between AOX and HMs concentration in the leachate. The other parameters of correlation results are given in
Appendix A Table A3. The dependence between the presence of organic substances in leachates and the content of HMs corroborated conclusions published by Tchounwou et al. [
45], and Przydatek et al. [
46], which point out the possible binding of HMs to organic substances, the reason being the high density and atomic mass of HMs. Thus, it can be assumed that, should an increased amount of organic substances be detected in the adjacent drill holes reaching into groundwater, this would be accompanied also by increased concentrations of HMs. This, however, did not happen; OP indicators were at minimum levels for all drill holes and the well (
Figure 5;
Figure 6) as well as the concentrations of HMs (
Figure 4) in the monitoring points for the period 2008–2013.
The analyzed relationships between the individual parameters suggest that a positive correlation existed between the content of organic substances (P
tot and N
tot) in the analyzed samples and the concentrations of most HMs (Cd, Zn, Cr, Pb, Ni, As). A similar positive correlation was observed between the COD, PAH, AOX and HMs. The given parameters, P
tot, N
tot and COD in particular, directly indicated the presence of organic substances in water [
38,
39]. The results from the LP indicate that, in addition to mineral substances, the landfill body also contained substances of organic nature to which HMs may potentially bind that subsequently can be washed out [
47]. However, the increased values (particularly those of N
tot and COD) were measured only in the LP. In the monitoring objects, the values of these parameters ranged at very low levels (
Figure 6). Hence, there was no groundwater contamination, nor any leakage of hazardous substances into the surrounding farmlands.
3.4. Phytotoxicity Parameters of Leachate Water
The results of the GI by leachates are presented in
Appendix A Figure A3. In the experimental period, leachates inhibited the model plant root growth at all tested rates. The average root GI in the model plant was 62.72%, 91.50%, 98.38%, and 99.76% for the leachate rates of 25%, 50%, 75%, and 90%, respectively. The measured data indicate that the increasing rate of leachate in the test had an increasingly inhibitory effect on the growth of white mustard (
Sinapis alba L.) roots.
The results of the GI at the semichronic exposure of duckweed (
Lemna minor L.) in the period of monitoring are presented in
Appendix A Figure A4. In the experimental period, all tested leachate rates (10%, 20%, and 100%) showed GI because all average values of GI > 0 (40.26%, 83.01%, and 100%, respectively). The results indicate that the increasing rate of leachate in the test had an increasingly inhibitory effect on the growth of duckweed (
Lemna minor L.). It was found in both model plants that the increasing concentration of leachates in the environment (i.e., with irrigation) was increasing their phytotoxicity. Similar results were published by Cheng et al. [
50]. However, for example Guerrero-Rodríguez et al., [
51] argue that in testing the impact of leachates on soil environment toxicity and yield of
Phaseolus vulgaris L., a dependence was found between the phytotoxic effect and the leachate dilution. If the leachate was added to irrigation water, the yield of
Phaseolus vulgaris L. grains decreased; however, the lowest yield was recorded in variants with the lowest dilution. In addition to this, the effect of grain crops and maize irrigation with leachate water was studied, with no demonstrable conclusions for the grain crops in terms of phytotoxicity [
51] and with no effect observed in maize [
52]. As to the measured laboratory values, it can be stated that the leachates from the experimental landfill might represent a problem in the agricultural area. On the other hand, the impact of the leachates for example on grain crops [
51], which represent the main agricultural crop in the studied area, has not yet been comprehensively characterized. Thus, the leachates can exhibit phytotoxicity in laboratory conditions, while their action in the field can be different, i.e., no adverse effect on cultivated plants. Their concentration in the soil environment can also be decisive [
50].
3.5. Possible Landfill Impact on the Agricultural Area
Literature reviews indicate that the impact on crops may result from the potential leaks through the base sealing layers into groundwater, surface runoff of rainwater and dust fall in dry periods [
14,
15]. Studies performed by Kaszubkiewicz et al. [
53] revealed that soils in the surroundings of legal landfills did not show any increased HMs content compared to the background. The presented study showed no HMs accumulations occurring in the soil–water environment near the landfill (
Figure 3). Moreover, Jahan et al. [
54] stated that if the landfill site is effectively managed, there is no objection to using the areas surrounding the landfill for agricultural production. Tracking the transport and fate of HMs near the landfill area is also important because soils can absorb leachate constituents. As a result, HMs absorbed by the soil can be easily taken up by plants. Subsequently this state is manifested by an effect on the growth and productivity of crops as well as on human and animal health [
55]. Equally worth emphasizing is the monitoring of ammonium concentration in the leachate and its impact on the surrounding agricultural areas [
54,
55,
56]. This is important because during their migration through the soil profile to deeper horizons they can contaminate underground sources of water [
56]. Since agricultural areas are permanently exposed to N pollution due to the widespread use of N fertilizers in agricultural practice [
54,
55,
56,
57,
58], attention should be given to minimizing the load of N compounds from external sources, including landfills and WM facilities.
The effect of leachate on higher plants (
Sinapis alba L. and
Cannabis sativa L.) was studied by Zloch et al. [
14,
18]. Phytotoxicity tests of concentrated leachate confirmed a GI higher than 90% (
Sinapis alba L.). A GI between 21.1% and 89.8% (depending on precipitation and month of sampling) was found when testing the leachate using
Cannabis sativa L. The tests clearly showed that susceptibility of higher plants to noxious agents is high and wastewater leakage into the environment might have fatal health consequences.
The surveys of groundwater quality showed no direct impact of pollution currently affecting groundwater quality (
Figure 1 and
Appendix A Figure A1). An indirect impact may be landfill dust fall on the surrounding agricultural crops. The actual impact on vegetation and its composition can be ascertained after tests, sampling and testing of groundwater samples from piezometers installed at a greater distance from the landfill. In setting up the installation distances of piezometers for groundwater monitoring, both groundwater flow directions and soil permeability (runoff lines on arable land surrounding the landfill are illustrated in
Appendix A Figure A1) should be taken into account. The potential impact of landfills located in agricultural areas can be exactly determined only through the extended scope of monitoring, including vegetation surveys.