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BY-NC-ND 3.0 license Open Access Published by De Gruyter Open Access July 26, 2016

Geostatistical study of spatial correlations of lead and zinc concentration in urban reservoir. Study case Czerniakowskie Lake, Warsaw, Poland

  • Piotr Fabijańczyk EMAIL logo , Jarosław Zawadzki and Małgorzata Wojtkowska
From the journal Open Geosciences

Abstract

The article presents detailed geostatistical analysis of spatial distribution of lead and zinc concentration in water, suspension and bottom sediments of large, urban lake exposed to intensive anthropogenic pressure within a large city. Systematic chemical measurements were performed at eleven cross-sections located along Czerniakowskie Lake, the largest lake in Warsaw, the capital of Poland. During the summer, the lake is used as a public bathing area, therefore, to better evaluate human impacts, field measurements were carried out in high-use seasons. It was found that the spatial distributions of aqueous lead and zinc differ during the summer and autumn. In summer several Pb and Zn hot-spots were observed, while during autumn spatial distributions of Pb and Zn were rather homogenous throughout the entire lake. Large seasonal differences in spatial distributions of Pb and Zn were found in bottom sediments. Autumn concentrations of both heavy metals were ten times higher in comparison with summer values.

Clear cross-correlations of Pb and Zn concentrations in water, suspension and bottom sediments suggest that both Pb and Zn came to Czerniakowskie Lake from the same source.

1 Introduction

An assessment of water and bottom sediments of lakes contaminated with heavy metals is of considerable importance due to negative effects of these metals on both humans and on the ecosystem of a lake. Natural geochemical background is characterized by low concentrations of heavy metals, from a fraction of a microgram to several hundred micrograms per liter of water [1, 2]. Heavy metals find their way into lakes due to human activities, as anthropogenic contamination, which can result in dramatic increases in their concentrations [3, 4]. The most common sources of heavy metal pollution are industrial activity, like coal power plants, steelworks or cement factories; through the use of fertilizers; and vehicle emissions. The degree of heavy metal contamination depends on urban development in the vicinity of a particular water body. Heavy metals get into water as a result of storm runoff flow and as a result of rainfall washing heavy metals from dust in the air [5]. Some studies have shown that particles originating from traffic are easily flushed from the ground surface by storm runoff [6]. It was found that surface runoff of storm runoff can contain over 120 μg/l of Zn and 20 μg/l of Pb [7]. Such high loads of heavy metals can pose a serious threat to the quality of urban lakes, because the water quality can rapidly exceed maximum allowable levels of heavy metals, even for recreational purposes.

Water and soil environments located in urban areas are subject to strong and long-lasting anthropogenic pressure[812]. Areas directly adjacent to urban lakes are often intensively managed i.e. by regulating their embankments or building urban infrastructure in the direct vicinity. Such processes combined with increased flow of organic pollutants can influence hydrographic conditions that in some cases can lead to eutrophication or even to drying up of the water bodies. For this reason, it is necessary to carry out permanent monitoring of the urban environment, taking into account the prevalence of selected heavy metals in waters and bottom sediments. Heavy metals accumulate in a superficial layer of the bottom sediments, and their total concentration changes considerably. This is typical of shallow urban reservoirs, which are typically characterized by significant variability of physico-chemical parameters of water [1315].

Sediments are an important sink for various pollutants such as heavy metals. Sediments, which provide habitats for many aquatic organisms, are commonly polluted with various kinds of hazardous and toxic substances, including heavy metals. They are considered to be one of the reservoirs for heavy metals accumulation in the hydrological cycle [1618] and can provide information about the impact of pollution sources [19, 20]. Waste materials, terrestrial runoff, disposal of liquid effluents from numerous urban, industrial, and agricultural activities can all accumulate in sediments. The majority of heavy metal emissions from anthropogenic sources accumulate in river and lake sediments [16, 21]. Discharge from sources such as smelters (Cu, Pb, Ni), metal based industries (Zn, Cr, Cd from electroplating), paint and dry preparations (Cd, Cr, Cu, Pb, Zn, Hg and Se), and petroleum refineries (As, Pb) can lead to metal accumulation in sediments [22, 23]. The pollution of the aquatic environment with heavy metals caused by urbanization and industrial development has become a major problem in recent years. Metals may be present in the lakes as dissolved species. Additionally, many metals become adsorbed to particulates or co-precipitated with carbonates and oxy-hydroxides, sulphides, and clay minerals [24]. Finally deposited metals in sediments occur in different geochemical forms, which have distinct mobility, biological toxicity, and chemical behaviors [18]. In the past years, tremendous efforts have been made to characterize the fates of heavy metals loading and distribution in rivers and lakes systems [20]. Heavy metals are generally transported to the sea by rivers or by industrial discharge. Their concentrations are generally directly correlated to organic matter content, so that problems are exacerbated in coastal sediments where heavy metals accumulate near population centers, river outflows, and industrial outfalls [25].

Studies of bottom sediments conducted in Sweden by Lindstrom and Hakanson [26] showed a substantial variation in concentrations of heavy metals in lakes. Bottom sediments contained high quantities of Pb, Cd, Zn and Hg, with low concentrations observed for Cr and Ni. 10% to 90% of heavy metals transported to the lakes accumulated in bottom sediments, including most of the Hg (ranging between 0.15 μg/g and 1.61 μg/g) and Pb (ranging between 75 μg/g and 413 μg/g), and less of the Ni (ranging between Ni 32 μg/g and 53.1 μg/g) and Cr (ranging between 15.2 μg/g and 38 μg/g). Accumulation in sediments was mainly controlled by processes such as sedimentation, and morphometry of lakes.

The potential environmental risks of trace elements in sediments are associated with total content and their speciation. Heavy metal concentration in sediment depends upon the factors influencing its accumulation and migration. Factors like depositional environment, particle size of the sediment and organic matter content play vital roles in accumulation of metal contaminants. Speciation analysis enables the identification of chemical forms of metals in the sediment. This analysis allows for identification and determination of the chemical forms of metals in a sample. For analytical purposes, Tessier distinguished five fractions (ion exchangeable, carbonate, adsorptive, organic and residual forms), in which metals are deposited [27]. Sediments are an important place for the deposition of various pollutants, including heavy metals. Use of speciation analysis allows the assessment of metal binding forms that are characterized by a high mobility and bioavailability in lake sediments [28].

The major geochemical phase for Pb in these sediments is the Fe-Mn oxide and residual phase. The speciation analysis also demonstrated that the fraction of hydrated iron – manganese oxides and residue is a dominant form of zinc occurrence. Easy-to-decompose compounds, which quickly release metals bonded in and to them into the environment, are another form of metal occurrence, which is hazardous to ecosystems. A lower percentage of the total Pb and Zn is bound to exchangeable-labile and carbonates phases [18, 29, 30].

Heavy metals deposited in mobile fractions are desorbed and can be transported in their mobile form for long distances under the influence of variable physical and chemical composition of water (salinity, dissolved oxygen, lowered pH) [15]. The phenomenon of heavy metals accumulation in bottom sediments [31, 32] constitutes a serious threat for the whole aquatic ecosystem and for people using this environment for recreational purposes. There can also be a serious threat to the environment adjacent to a water body because sediment-bound heavy metals can be transported during floods [33].

Studies conducted so far indicated that the decrease in water quality of urban lakes in Warsaw resulted mostly from the presence of phosphates or nitrates that has caused strong eutrophication of the lake. Low water quality was also caused by concentrations of heavy metals, particularly Zn, Cu, Cd and Pb. The continuous migration of heavy metals to the lake caused their accumulation in the bottom sediments and then the periodic release of those metals into the water. The result of such processes can be the degradation of biological life in the lake and its surroundings [13].

The aim of this work was to investigate the levels and spatial distribution of pollution of water and sediments of a lake located within a large urban agglomeration by lead (Pb) and zinc (Zn), as these heavy metals are common pollutants related to traffic and industrial activities. Additionally, spatial correlations and cross-correlations of aqueous, suspension and sediment bound Pb and Zn were analyzed.

2 Material and methods

2.1 Study area

Czerniakowskie Lake is the largest natural lake located almost in the center of the Warsaw agglomeration. This lake is a part of the former river-bed of the Vistula River which is known as the Praski terrace. It emerged as a relic on a flood terrace of the Vistula River during the Neolithic age, that is about 6 500 years ago. Since 1987 the lake has been part of a water-landscape reserve comprising an area of 45 000 m2, while the lake itself covers an area of 20 000 m2. The lake constitutes a part of a large hydrological system, connected into one water network. At present Czerniakowskie Lake is also used for recreation. There are numerous bathing beaches and recreation grounds for local people, located mostly near the middle of the southern part of the lake. In the direct proximity of the lake (Fig. 1) there are allotments that can also influence the quality of water in lake. One of the most important factors influencing the quality of the lake and its surroundings is the neighboring Siekierki power plant, which, with its capacity of 622MW, is one of the largest power plants in Poland. Almost half of the eastern lakeside is located in direct vicinity of nearby residential allotments, which are potential sources of pollutants such as nitrates, phosphates and heavy metals.

Figure 1 Study area of Czerniakowskie Lake and its vicinity.
Figure 1

Study area of Czerniakowskie Lake and its vicinity.

2.2 Measurements of heavy metals concentration

The measurement network was formed from 10 crosssections located across the lake, with 3 sample points in each cross-section (Fig. 1). The average distance between the cross-sections was about 250 m. At each sample point, measurements were carried out in water, suspension and bottom sediment. All measurements were performed twice, once during the summer and once in the autumn.

2.2.1 Water measurements

A Ruttner sampler, manufactured by KC Denmark Research Equipment, Denmark, was used to collect water samples from the near surface layer (between, 0 – 0.5 m in lake depth). Sampled water was filtered with the use of a Fison filtering system, applying nitrocellulose membrane filters with the pore diameter 0.45 μm. Filtered water was mineralized by placing an aliquot in an acid mixture (5 ml HNO3, 2 ml HClO4 and 1 ml HF). All reagents were analytically pure (Analytical Research Grade, AR). Solutions were prepared with the use of double distilled deionized water. Graphite Furnace Atomic Absorption Spectrometry (GF-AAS) was used to determine aqueous Pb and Zn concentrations.

Table 1

Descriptive statistics of concentration of aqueous Pb and Zn, suspension bound Pb and Zn, sediment bound Pb and Zn measured in Czerniakowskie Lake in summer (S) and autumn (F).

Water [mg/l]Suspension [mg/l]Sediments [mg/kg dry mass]
SummerAutumnSummerAutumnSummerAutumn
ZnAverage0.0640.0540.0820.072109.084176.929
Median0.0490.0360.0670.07063.106197.905
Standard deviation0.0530.0390.0450.020162.526138.441
Minimum0.0100.0130.0350.03517.9529.207
Maximum0.2400.1400.1700.122692.022456.188
SummerAutumnSummerAutumnSummerAutumn
PbAverage0.0960.0940.2040.02418.26136.303
Median0.0830.0840.1990.02318.61042.660
Standard deviation0.0550.0630.0350.01211.97424.030
Minimum0.0190.0100.1370.0033.0001.500
Maximum0.2060.2370.2640.04842.79075.300

2.2.2 Suspension measurements

Measurements of suspension-bound Zn and Pb were performed using solid phase residue left on nitrocellulose 0.45 μm membrane filters left after filtering the water samples. A membrane filter with retained solids was dried to constant weight at 105°C, and subsequently underwent mineralization. Wet mineralization was conducted by placing an aliquot in an acid mixture (5 ml HNO3, 2 ml HClO4 and 1 ml HF). All reagents were analytically pure (Analytical Research Grade, AR). Solutions were prepared with the use of double distilled deionized water. Suspension-bound Zn and Pb concentrations were determined using Flame Atomic Adsorption Spectroscopy (FAAS), based on reference curves determined for a series of already-prepared MERCK reference solutions. Limits of Detection (LODs) for particular metals were calculated 3fold for two standard deviations: Pb (0.05), Zn (0.01). In order to validate the analysis and precision of measurement, mineralization was conducted on a reference material, (CRM) Till-3, with a pre-determined metal content. Recovery rate of particular metals from reference sediments oscillated in the range 93÷110%. Error on comparison of the test results did not exceed 10%.

2.2.3 Sediment measurements

10 cm thick samples were collected from the surface layer of the sediments using a sampler, Kajak type KC Denmark Research Equipment, Denmark. The collected samples were packed into polyethylene bottles and transferred to the laboratory, where they were dried at room temperature. Dry sediment samples were used to analyze grain size fractions and Pb and Zn concentrations. In order to determine the relationship between grain size and metal content, the sediment samples were fractionated into six grain sizes by a controlled sieve shaker. Only sediment samples with a grain size lower than 100 μm were used, because this fraction was the dominant fraction of the bottom sediment and did not contain any accidental pollution. 1g of dried sample was digested with a mixture of HNO3/HClO4 (3:1) acids. Mineralization was conducted in a digestion block, with a bottom sediment and acid mixture placed in a Teflon container. The concentration of Pb and Zn was determined using atomic absorption spectrometry (FAAS Philips PU9100X), based on reference curves determined for a series of already-prepared MERCK reference solutions. Limits of Detection (LODs) for particular metals were calculated 3-fold for two standard deviations: Pb (0.05), Zn (0.01). In order to validate the analysis and precision of measurement, mineralization was conducted on a reference material, (CRM) Till-3, with a pre-determined metal content. Recovery rate of particular metals from reference sediments oscillated in the range 93÷110%. Error on comparison of the test results did not exceed 10%.

Quality assurance and quality control (QA/QC) for metals in sediment samples were estimated by determining metal concentrations in the Merck Standard solutions (Merck, Darmstadt, Germany). The detection limit was calculated based on the estimated instrumental detection limit assuming that 1 g of a sample is digested or diluted to 100 ml. Detection limits (mg/kg of dry matter) for Pb and Zn were 0.003 and 0.001, respectively.

2.3 Calculation methods

Because the number of samples was limited to 30, spatial distributions of Pb and Zn concentration were calculated using the Inverse Distance Weighting method (IDW). All calculations concerning spatial distributions were calculated using the ArcGIS program.

Analyses of spatial correlations of Pb and Zn concentration were performed using semivariance and experimental variograms [34]. A variogram is defined as a half of variance of difference of random variable Z in position x and x + h, where vector his a lag vector and vector xcovers all possible positions of the random variable.

γ(h)=12Var{Z(X)Z(X+h)}(1)

Additionally, cross-correlations between measured concentrations of Pb and Zn were analyzed using experimental cross-variograms. In this case it is possible to analyze the spatial correlations similarly to a variogram, but between two different types of measurements or measured indicators, denoted by Z and S:

γ(h)=12Var{Z(X)S(X+h)}(2)

After the experimental variograms and cross-variograms were calculated they were modeled using the spherical model. Parameters of variogram models were analyzed individually, namely a nugget effect, a range of correlation, and a sill. A nugget effect is a value of semivariance for a distance vector hequal to zero, and can be related to the spatial variability for distances shorter than the distance between samples and to possible measurement errors. A range of correlation is a distance where the variogram flattens and achieves a sill. The lower the nugget effect, and especially the nugget-to-sill ratio, then the analyzed phenomenon is more spatially correlated.

3 Results and discussion

3.1 Aqueous Pb and Zn

Spatial distributions of the aqueous Pb and Zn were comparable in summer and autumn, as shown in spatial distributions in Fig. 2 and 3. The maximum observed concentration of aqueous Zn was 0.24 mg/l in the summer and 0.14 in the autumn. In the case of aqueous Pb the maximum observed concentrations for both seasons were equal to about 0.2 mg/l.

Figure 2 Spatial distributions of concentration of aqueous Zn, suspension bound Zn and sediment bound Zn in Czerniakowskie Lake as measured in summer (S) and autumn (F).
Figure 2

Spatial distributions of concentration of aqueous Zn, suspension bound Zn and sediment bound Zn in Czerniakowskie Lake as measured in summer (S) and autumn (F).

In both seasons, concentrations of aqueous Pb and Zn were highest in the vicinity of the bridge, which was located in the center part of the lake. These concentrations were probably caused by the traffic dust that was deposited on the surface of the bridge and later flushed into the lake by storm runoff. The hot-spot of aqueous Zn and Pb that was observed in the vicinity of the bridge had much higher values in summer, reaching up to 0.24 mg/l and 0.20 mg/l for Pb and Zn, respectively. These concentrations may have been caused by the fact that in Warsaw the average rainfall in the summer can be up to 60% higher than that in the autumn.

3.2 Suspension bound Pb and Zn

The spatial distribution of the suspension bound Zn measured in the summer was characterized by two distinctive hot-spots. The first one was observed in the vicinity of the bridge across the lake and the second one in that part of the lake where the most recreational beaches were located (Fig. 2). These hot-spot values of suspension bound Zn reached 0.17 mg/l. In the rest of the lake the measured concentrations of suspension bound Zn were lower, especially in the central part of the lake, where it did not exceeded 0.06 mg/l. In the autumn, the spatial distribution of suspension bound Zn was homogenous across the entire lake, with the dominant concentration of suspension bound Zn was ranging between 0.06 and 0.09 mg/l (Fig. 2).

The spatial distribution of suspension bound Pb measured in the summer was characterized by a hot-spot of about 0.26 mg/l in the vicinity of the bridge, and additional large hot-spot in the northern end of the Czerniakowskie Lake (Fig. 3). In the autumn, measured suspension bound Pb concentrations were considerably lower, up to ten times in some parts of the lake. This may have been caused by quick migration of the suspension bound Pb to the sediment phase in the summer, and as a consequence, in autumn lower suspension bound concentrations were observed.

Figure 3 Spatial distributions of concentration of aqueous Pb, suspension bound Pb and sediment bound Pb in Czerniakowskie Lake as measured in summer (S) and autumn (F).
Figure 3

Spatial distributions of concentration of aqueous Pb, suspension bound Pb and sediment bound Pb in Czerniakowskie Lake as measured in summer (S) and autumn (F).

3.3 Sediment bound Pb and Zn

During the summer sampling periods the concentrations of aqueous Pb and Zn were mostly observed as hot-spots, while in autumn these concentrations were homogenous across the entire lake. Such observations suggested that a large quantity of the Zn and Pb was transported into the water of Czerniakowskie Lake during summer, and after that Zn and Pb were gradually accumulated in the suspension and bottom sediments. The effect of this accumulation was observed in the autumn as heightened concentrations of suspension and sediment bound Zn and Pb, but was simultaneously spatially homogeneous across the entire lake.

In the autumn the concentrations of sediment bound Zn was up to five times higher than the concentrations observed in the summer (Fig. 2). Similarly, also the concentrations of Pb in sediments were higher than those observed in the autumn, but however, this difference was not as significant as in the case of Zn (Fig. 3).

3.4 Spatial correlation and cross-correlation of Pb and Zn

Aqueous, suspension and sediment bound Pb did not show strong spatial correlations. Variograms presented in the Fig. 4a, 4b, 4c were characterized as pure nugget effect.

Figure 4 Standardized experimental variograms of concentration of aqueous Pb and Zn, suspension bound Pb and Zn, sediment bound Pb and Zn.
Figure 4

Standardized experimental variograms of concentration of aqueous Pb and Zn, suspension bound Pb and Zn, sediment bound Pb and Zn.

Figure 5 Experimental cross-variogram and its model between aqueous Pb and Zn, measured in summer.
Figure 5

Experimental cross-variogram and its model between aqueous Pb and Zn, measured in summer.

Figure 6 Experimental cross-variogram and its model between aqueous Pb and Zn, measured in autumn.
Figure 6

Experimental cross-variogram and its model between aqueous Pb and Zn, measured in autumn.

The model of the variogram for aqueous Zn was characterized by two overlapping ranges of spatial correlations, as presented in 4b. The first range of correlation was equal to approximately 225 m, and was probably related to the variability of aqueous Zn within a single cross-section of the measuring network. The second range of spatial correlation, visible in the 4b as an increase of the semivariogram between 300 m and 450 m, corresponded roughly to the distance between different cross-sections. The partial sill of the first structure was equal to about 0.5, with the second structure equal to 0.8 (Fig. 4b). This observation suggests that spatial variability of aqueous Zn within a cross-section was slightly lower than the variability across the entire lake. Suspension and sediment bound Zn was characterized by no visible spatial correlations, and the variograms (Fig. 4d and 4f) only exhibit a pure nugget effect.

Cross-correlations were only investigated for aqueous Pb and Zn, separately for the summer and autumn. Both cross-variograms were modeled by using similar spherical models. The range of correlation was equal to about 500m (Table 2). Both cross-variograms were also characterized by zero nugget effect. Because the spatial correlations of aqueous Pb and Zn during the summer and autumn were characterized by similar models of spatial correlations, it can be concluded that spatial cross-correlations did not change with time.

Table 2

Parameters of variogram and cross-variogram models of concentration of aqueous Pb and Zn, suspension bound Pb and Zn, sediment bound Pb and Zn.

Standardized variograms
Range

[m]
Nugget effect

[-]
Sill

[-]
Aqueous Pb-1.0 (pure nugget)-
Suspension bound Pb-1.0 (pure nugget)-
Sediment bound Pb-1.0 (pure nugget)-
Aqueous Zn2250.520.58
Suspension bound Zn-1.0 (pure nugget)-
Sediment bound Zn-1.0 (pure nugget)-
Cross-variograms
Range

[m]
Nugget effect

[(mg/l)2]
Sill

[(mg/l)2]
Aqueous Pb-Zn, summer500-0.002
Aqueous Pb-Zn, autumn500-0.0005

Cross-variograms of the suspension and sediment bound Pb and Zn were not modeled because of poor spatial correlations.

Using variograms it was not possible to detect spatial variability of Zn and Pb that were transported into the lake, and only the cross-variograms of aqueous Zn and Pb showed weak spatial correlations. During the autumn spatial distributions of the aqueous, suspension and sediment bound Zn and Pb were homogenous across the entire Czerniakowskie Lake. This suggests that in this shallow lake, which has a very shallow depth in comparison with the lake’s length, mixing processes were substantial, which yielded poor spatial correlations. Aqueous, suspension and sediment bound Zn and Pb were mostly characterized by pure nugget effect, which suggests that short-scale variations were predominant, as no long-distance spatial correlations were observed. The short-scale variations of the aqueous, suspension and sediment bound Zn and Pb were mostly governed by the processes of adsorption, desorption and variable mobility and bioavailability of the heavy metals. As a consequence, these variations could not be described by the spatial correlation models.

Measured results showed that at Czerniakowskie Lake the major anthropogenic pressure were related to dusts containing heavy metals that were flushed via storm runoff from the bridge that spans across the lake. This suggests that the proper storm runoff management is needed, because without it, high concentrations of heavy metals may be gradually accumulated in the sediment phase. As a consequence, the Czerniakowskie Lake may become unsuitable for recreational purposes.

4 Conclusions

Calculated spatial distributions of lead and zinc concentrations in water differ in summer and autumn. In the summer several hot-spots of both Pb and Zn were observed mostly in the vicinity of the most likely pollution source. In autumn the hot-spots of concentration of Pb and Zn were less pronounced, and Pb and Zn concentrations were spatially more homogenous across the entire lake.

The most substantial spatial differences were found between concentration of Pb and Zn in bottom sediments. In autumn the concentrations of both studied heavy metals in bottom sediments were considerably higher in comparison with their concentrations in summer.

It was possible to determine characteristic ranges of spatial correlations, especially cross-correlations between Pb and Zn concentration in water. There was a noted increase in the range of correlation between bottom sediments, suspension and water, with water showing the greatest spatial correlations.

Analysis of cross-correlations between Pb and Zn concentration in water showed that there was no change in the range of correlation between summer and autumn. The nugget effect also did not change between the two seasons. This analysis of cross-correlations can be used to determine if there are differences between the ways two indicators are spreading in time in water.


Piotr Fabijańczyk: Faculty of Building Services, Hydro and Environmental Engineering, Warsaw University of Technology, Warsaw, Poland; Tel.: 048 022-825-18-63; Fax: 048 022-234-5410

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Received: 2014-10-10
Accepted: 2015-10-9
Published Online: 2016-7-26
Published in Print: 2016-7-1

© 2016 P. Fabijańczyk et al., published by De Gruyter Open

This work is licensed under the Creative Commons Attribution-NonCommercial-NoDerivatives 3.0 License.

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