*Article* **Long-Term Leaching Behavior of Organic and Inorganic Pollutants after Wet Processing of Solid Waste Materials**

**Maria Prieto-Espinoza 1,2 , Bernd Susset <sup>1</sup> and Peter Grathwohl 1, \***


**Abstract:** The recycling of mineral materials is a sustainable and economical approach for reducing solid waste and saving primary resources. However, their reuse may pose potential risks of groundwater contamination, which may result from the leaching of organic and inorganic substances into water that percolates the solid waste. In this study, column leaching tests were used to investigate the short- and long-term leaching behavior of "salts", "metals", and organic pollutants such as PAHs and herbicides from different grain size fractions of construction & demolition waste (CDW) and railway ballast (RB) after a novel treatment process. Specifically, silt, sand and gravel fractions obtained after a sequential crushing, sieving, and washing process ("wet-processing") of very heterogeneous input materials are compared with respect to residual contamination, potentially limiting their recycling. Concentrations in solid fractions and aqueous leachate were evaluated according to threshold values for groundwater protection to identify relevant substances and to classify materials obtained for recycling purposes according to limit values. For that, the upcoming German recycling degree was applied for the first time. Very good agreement was observed between short and extensive column tests, demonstrating that concentrations at L/S 2 ratios are suitable for quality control of recycling materials. Different solutes showed a characteristic leaching behavior such as the rapid decrease in "salts", e.g., SO<sup>4</sup> <sup>2</sup><sup>−</sup> and Cl −, from all solid fractions, and a slower decrease in metals and PAHs in the sand and silt fractions. Only the gravel fraction, however, showed concentrations of potential pollutants low enough for an unlimited re-use as recycling material in open technical applications. Sand fractions may only be re-used as recycling material in isolated or semi-isolated scenarios. Leaching from heterogeneous input materials proved harder to predict for all compounds. Overall, column leaching tests proved useful for (i) initial characterization of the mineral recycling materials, and (ii) continuous internal (factory control) and external quality control within the upcoming German recycling decree. Results from such studies may be used to optimize the treatment of mixed solid waste since they provide rapid insight in residual pollution of material fractions and their leaching behavior.

**Keywords:** mineral recycling material; leaching test; heterogeneity; compliance testing

#### **1. Introduction**

The largest solid waste stream in Germany with an annual volume of more than 275 million tones comprises 32% of construction and demolition waste (CDW), of which about 90% is reused [1]. Recycling mineral waste has a lot of advantages in terms of sustainability and economical aspects. To increase recycling potential, more and more companies start to treat excavated soil-stone mixtures, and demolition waste or railway ballast, combining crushing with dry and wet sieving and washing processes (wet processing). However, the reuse of mineral materials may pose potential risks of environmental pollution, resulting from leaching of organic and inorganic substances into percolating water and, ultimately, into groundwater [2–5].

**Citation:** Prieto-Espinoza, M.; Susset, B.; Grathwohl, P. Long-Term Leaching Behavior of Organic and Inorganic Pollutants after Wet Processing of Solid Waste Materials. *Materials* **2022**, *15*, 858. https:// doi.org/10.3390/ma15030858

Academic Editor: Andrea Petrella

Received: 12 December 2021 Accepted: 19 January 2022 Published: 23 January 2022

**Publisher's Note:** MDPI stays neutral with regard to jurisdictional claims in published maps and institutional affiliations.

**Copyright:** © 2022 by the authors. Licensee MDPI, Basel, Switzerland. This article is an open access article distributed under the terms and conditions of the Creative Commons Attribution (CC BY) license (https:// creativecommons.org/licenses/by/ 4.0/).

In this context, the risk concerning potential contaminants in such materials must be addressed by their leaching potential into water rather than assessing total concentrations in the solid phase. Column leaching tests and batch shaking tests are frequently used to assess the transfer of contaminants into water [3,6,7]. Column leaching tests are preferred because they allow for assessing time-dependent behavior and simulating the flow of water through solid materials closer to natural conditions [8,9]. In Germany, the standard DIN 19528 (2009) is used for examining the leaching potential of inorganic and organic substances from solid materials [10]. Generally, "extensive column tests" are performed to characterize and evaluate the long-term leaching behavior of contaminants in which eluates are collected at different liquid-to-solid (L/S) ratios (e.g., 0.3, 1, 2, 4 and 10 L/kg). L/S ratios represent the time after which a certain volume of water has percolated through the solid material in the column (in L/kg dry matter). For compliance testing, "short column tests" may be employed that provide results of cumulative concentrations at one fixed L/S of 2 L/kg, which then are compared with threshold values set by special regulations [5,9,11]. Initially, equilibrium concentrations are often observed in column eluates [12]. The decrease in concentrations with increasing L/S ratios (or time) may be due to depletion of highly soluble substances, or a shift to non-equilibrium conditions because of mass transfer limitations (e.g., slow intraparticle diffusion) indicated by an extended tailing of the solute concentrations in the leachate [13]. The shift between equilibrium to non-equilibrium conditions may depend on initial conditions [12], flow velocities, grainsizes, sorption capacity and contaminant release kinetics [8,9,12]. Typically, three basic leaching scenarios can be described for (i) fast leaching substances such as "salts" (e.g., sodium, potassium, chloride), where a rapid decline in concentrations in column effluents is observed (at L/S < 2 L/kg); (ii) intermediate compounds such as some metals, where mass release is governed by leaching parameters such as pH, redox conditions, ionic strength and DOC-complexations [14–17], and; (iii) for strongly sorbing compounds such as PAHs, where equilibrium concentrations prevail over extended periods of time [3].

Column leaching tests are thereby proposed as a common procedure for the evaluation of environmental qualities of solid waste recycling materials [5,9,18]. In Germany, the upcoming recycling directive [19] is based on improved methods for groundwater risk assessment to derive a new regulatory framework for the reuse of solid waste materials. For a given substance, the concentration level avoiding any significant alteration of the chemical status of groundwater is defined as the "insignificance threshold" concentration ("GFS", in German "Geringfügigkeitsschwelle") [20]. The GFS values are based on ecoand human toxicological tests and are not intended to set a quality goal for groundwater, but rather reflect a groundwater status unaffected by human activity [20].

Concentration limits at L/S 2 eluates are set depending on the type of mineral recycling material (e.g., CDW, RB, steel slag etc.), the type of technical application (open technical applications and isolated or semi-isolated technical applications protected from seepage water), the distance to the groundwater table, and the soil characteristics of the underground [19]. If the quality of the recycling material shows high variability, different "material classes" are defined by different sets of limit concentrations in eluates from the same mineral recycling material, so-called "material values" (e.g., material class RC–1, highest quality) [11]. The comparison of concentrations in L/S 2 eluates with GFS values and/or limit values of "material classes" will ultimately define permissible applications of the mineral recycling material. This concept is implemented in Germany within the upcoming recycling directive [19] with a quality assurance system and material-specific testing programs, where the quality of mineral recycling materials is assessed based on extensive column percolation tests [10]. Furthermore, short-term column percolation tests [10] are performed for internal and continuous external quality control [11].

In this study, short and extensive column tests were performed to examine the leaching of organic and inorganic substances from railway ballast (RB), and construction and demolition waste (CDW), which both underwent a sophisticated washing and grain size separation process, so-called wet processing. The input material ('In', highly heterogeneous

RB or CDW) as well as its separated grain-size fractions such as silt ('U', <0.063 mm), sand ('S', 0–2 mm) and gravel ('G', 2–8 mm) were examined. The aim of this study was (i) to characterize comprehensively the leaching behavior of organic and inorganic contaminants from input and recycled material fractions (i.e., RB and CDW); (ii) to assess the quality of the different grain size fractions with respect to threshold values for potential risks of groundwater contamination and material values; (iii) to compare results of the short and extensive tests, and; (iv) to examine the long-term leaching behavior of the investigated substances. size fractions such as silt ('U', < 0. sand ('S', 0–2 mm) and gravel ('G', 2–

"material classes" are defined by different sets of limit concentrations in eluates from the

and/or limit values of "material classes" will ultimately define permissible applications of

called "material values" (e.g., material class RC–

called wet processing. The input material ('In', highly heterogene-

#### **2. Materials and Methods**

#### *2.1. Materials*

In total, 3 sets of construction and demolition waste (CDW1, CDW2 and CDW3) and 1 set of railway ballast (RB) were examined to reflect different sources of recycling materials and their variability. Samples were collected in December 2017 (RB and CDW1), March 2018 (CDW2) and May 2018 (CDW3). The original CDW material was a mixture of soil, demolition and construction waste. At the recycling plant, input materials were crushed and "cleaned" in a complex washing process. All solid waste materials were separated into different grain size fractions (Figure 1). Large fragments were sieved into different gravel-size fractions: 32–50 mm, 16–32 mm, 8–16 mm and 2–8 mm. Smaller particles in suspension underwent a centrifugation process, wherein the sand fraction (0–2 mm) was separated and further washed. Finally, the silt fraction (<0.063 mm) was separated from the suspension by centrifugation with the addition of flocculants and polymers. The input material as well as the silt, sand and gravel (2–8 mm) fractions were used for the column tests without further sieving or crushing (Figure 1). In addition, the aqueous solution (referred to as "washing water") used for separation and cleaning of the solid fractions at the recycling plant was analyzed. crushed and "cleaned" in a complex washing process. All solid waste materials were sep- – – – – – – as "washing water") used for separation an

**Figure 1.** Solid waste fractions of construction and **Figure 1.** Solid waste fractions of construction and demolition waste (CDW3) as received from the recycling plant: input material (In) as well as silt (U), sand (S) and gravel (G) fractions obtained after wet-treatment.

#### *2.2. Experimental Setup*

Prior to the column leaching tests, the gravimetric water content (*w*) of the solid fractions was determined by weighting and drying the wet material in an oven for 24 h at 105 ◦C. The dried material was further used to determine its volume and grain density using a gas pycnometer (micromeritrics/AccuPyc 1330). Quartz sand was used as a reference standard material (density: 2.65 g/cm<sup>3</sup> ). The final values were set by measuring 10 times the same material until reaching a standard deviation of less than 0.005 g/cm<sup>3</sup> . In order to increase the permeability and to prevent mobilization of fine particles, the input material and silt fraction were mixed with clean quartz sand, as suggested elsewhere (e.g., [5]). From the input material, only particle fractions smaller than 32 mm were used (see Figure 1).

Column leaching tests were carried out according to the German standard DIN 19528 (2009) in a dark laboratory at a constant temperature of 20 ◦C [10]. The DIN 19528 has

been validated for investigations on long-term leaching of salts and heavy metals from incineration bottom ash [9,21], comparisons with batch and lysimeters tests [3], and the effect of contact time in column percolation tests [9]. Furthermore, Lin et al. [5] recently proposed an optimization of the short column percolation tests (at L/S 2 eluates; DIN 19528,) by approving the use of sand admixtures in coarse grain fractions.

In total, 27 short column tests were performed for all fractions of the solid waste (i.e., RB, CDW1, CDW2 and CDW3), including 3 controls containing only a 3 cm layer of quartz sand. Short column eluates were collected until a L/S of 2 L/kg and analyzed for salts, metals, and organic substances such as BTEX, PCBs, herbicides and PAHs. In addition, 6 extensive column leaching tests were performed for sample CDW3 to examine the long-term leaching behavior of contaminants at L/S of 0.1, 0.3, 1, 2, 4 and 10 L/kg, including 1 control column. For the silt fraction, the earliest column eluate was collected at L/S 0.5 ratio. Glass columns with an inner diameter of 5 cm and a length of 30 cm were used for the sand and gravel fractions, whereas glass columns with an inner diameter of 7 cm were used for the input material and the silt fraction previously mixed with quartz sand. Before packing the samples into the columns, a 1 cm layer of quartz sand was placed at the bottom for better distribution of the water flow through the column inlet. A second quartz sand layer was placed at the top, at a filling height of about 28 cm, to prevent the release of fine particles. Additionally, glass wool was placed at the inlet and outlet openings. Teflon tubes were connected to the column inlets and the clean water reservoir consisting of a 50 L glass bottle containing Milli-Q water. The flow rate was set using a peristaltic pump (IPC 8, ISMATEC), and adjusted to allow a contact time of 5 h during the leaching tests. The initial flooding of the columns with clean water lasted approximately 2 h. Column eluates were collected in amber glass bottles at the corresponding L/S ratios and stored at a temperature of 20 ◦C until further analysis. Given that biodegradation and volatilization of organic compounds can occur, columns for PAHs were run in parallel—one for the analysis of ions and metals and the other for PAHs only. For PAH analysis, the collecting bottles previously contained 10 mL of cyclohexane (to avoid any biodegradation during sampling and storage), and an internal standard (10 µL, 5 perdeuterated PAHs according to DIN 38407–39 in toluene, each perdeuterated PAH 20 ng/µL).

#### *2.3. Analytics*

2.3.1. Turbidity, Electrical Conductivity, Ion Chromatography and DOC

All column eluates were analyzed for turbidity, electrical conductivity (EC, HACH LANGE), and pH (inoLab® pH 7110, WTW) within the first 2 h after collection. After filtration at 0.45 µm, major ions were analyzed by ion chromatography (DIONEX, DX-120). Total Organic Carbon (TOC) and Dissolved Organic Carbon (DOC) were measured via a TOC analyzer (Elementar, Vario TOC).

#### 2.3.2. Metals, Phenols, EOX, PCBs, PHCs, Cyanide and Herbicides

Solid concentrations of heavy metals, petroleum hydrocarbons (PHCs, C10-C40), polychlorinated biphenyls (PCB), phenols, extractable organic halides (EOX) and cyanide were analyzed at the Gewerbliches Institut für Umweltanalytik GmbH (Industrial Institute for Environmental Analysis, Teningen, Germany). PHCs, PCBs, phenols, EOX and cyanide concentrations were measured by gas chromatography−tandem mass spectrometry (GC-MS/MS, Agilent).

Aqueous column eluates (aliquots of 20 mL) were filtered at 0.45 µm and acidified (HNO3) prior to the analysis of heavy metals via inductively coupled plasma mass spectrometry (ICP-MS, Agilent). The herbicides atrazine, simazine, bromacil, desethylatrazine, hexazinone, dimefuron, diuron, flumioxazin, thiazafluron and ethidimuron were measured in 20 mL aliquots of column eluates by liquid chromatography−tandem mass spectrometry (LC-MS/MS, Agilent). All substances were measured according to protocols described in [20].

#### 2.3.3. Polycyclic Aromatic Hydrocarbons (PAHs)

PAHs were measured in both solids and column eluates using GC-MS (Agilent/HP 5973). For solid concentrations, PAHs were extracted by Accelerated Solvent Extraction (ASE 300 DIONEX, Thermo Scientific), a technique that utilizes organic solvents at high temperature and pressures. Approximately 40 g of the solid samples were placed in the sample cell, along with 47 mm diameter filters on both ends of the extraction cell. Samples were extracted sequentially first with acetone and then with toluene (50 mL extracts) at a pressure of 100 bars and 100 ◦C [22]. Aqueous column eluates were extracted by liquid-liquid extraction. The bottles containing the column eluates along with 10 mL of cyclohexane (CH) and 10 µL of internal standard (10 µL, 5 perdeuterated PAHs according to DIN 38407–39, in toluene, each perdeuterated PAH 20 ng/µL) were horizontally shaken for 1 h (at 150 rpm), and subsequently filled with Milli-Q water until the solvent reached the bottleneck. The bottles were left overnight, and cyclohexane extracts were retrieved and treated with anhydrous sodium sulfate. All extracts were reduced to 200 µL by means of a nitrogen flow.

#### **3. Results and Discussion**

#### *3.1. Pollutant Screening in Solid Fractions*

Prior to the column tests, solids were analyzed for determination of initial concentrations and characterization of the materials according to precautionary values for soils [19]. Concentrations of PCBs were present in some solid fractions, but did not exceed precautionary values for soils (Appendix A, Table A1) [19]. Phenols, cyanide and EOX were not detected in the solid samples, except for the input material and the sand fraction of CDW2 with EOX concentrations of <0.5 µg/kg. PHCs (C10-C40) were present in concentrations below the limit value of material class BM-0\* (<300 mg/kg; Table A1). For EOX and PHCs no precautionary values exist for soils [19]; therefore, limit values for material classes are used (Table A1). Metals exceeding precautionary values were detected in silt fractions of both RB and CDW materials, e.g., As (>10 mg/kg), Pb (>40 mg/kg), Cu (>20 mg/kg) and Zn (>60 mg/kg), while both, the silt and sand fractions exceeded the limit of solid concentrations for Cd (>0.4 mg/kg), Cr (>30 mg/kg) and Ni (>15 mg/kg) (Figure 2 and Table A1). Further metal solid concentrations are given in Table A1 for information.

Globally, the silt fraction was the most contaminated particularly with Cr and Cu in RB, and with the 16 PAHs in CDW samples (>3 mg/kg; Figure 2). The variability of solid concentrations in the different CDW samples is low for metals but high for PAHs. Moreover, RB shows different solid concentration patterns than CDW. The high variability of solid concentrations demonstrates that solid waste materials should preferably be examined as individual samples and according to grain size for intended use prior to recycling applications. While the washing process of the solid material into different grain size fractions should be considered as an important step for the separation of fractions suitable for recycling applications, the treatment is obviously not sufficient to clean up the materials to reach precautionary values for the sand and silt fractions. Only the gravel fraction reached concentrations below precautionary values (Figure 2), with the exception of PAHs in the gravel fraction of CDW3. The solid concentrations of PAHs in the sand and silt fraction of CDW 2 exceed the material value of RC-3 (20 mg/kg), and based on this, it cannot be reused and would have to be landfilled.

**Figure 2.** Concentrations of (**a**) As, (**b**) Cu, (**c**) Ni, and (**d**) the sum of the 16PAHs in solids: Input material (In), silt (U), sand (S) and gravel (G) of railway ballast (RB) and three sets of construction and demolition waste (CDW); input represents the material prior to separation into the different fractions (silt, sand and gravel), and the red dashed lines indicate precautionary values for sandy soils [19]. Error bars represent uncertainties in measurements.

#### *3.2. Contaminant Concentrations in Eluates of Short Leaching Tests and Washing Water*

− − − Column eluates were examined at L/S 2 of the recycling materials RB, CDW1, CDW2 and CDW3 from 4 different solid fractions: input material (In), silt (U), sand (S) and gravel (G); column parameters are listed in Table 1. Further detailed concentrations are listed in the Supplementary Materials. The gravel fractions of both RB and CDW materials showed the lowest concentrations in the leachates, without exceeding GFS values, except for V (> 4 µg/L) and herbicides in RB (Appendix A, Table A2 and Supplementary Materials). Herbicides were only detected in RB (Table A2), where sand was the most contaminated fraction exceeding GFS values (>0.1 µg/L per herbicide). In general, leachates from CDW materials showed higher concentrations than those from RB, except for As and Mo (Figure 3). SO<sup>4</sup> <sup>2</sup><sup>−</sup> showed the highest concentrations up to 270 mg/L, followed by Cl − and NO<sup>3</sup> <sup>2</sup>−. Furthermore, the highest eluate concentrations at L/S 2 were observed in silt fractions of CDW, followed by the input material, and sand and gravel fractions (Figure 3). Overall, the gravel fractions of RB and CDW materials proved to be the least contaminated, and thus suitable for a free re-use as recycling material in open technical applications.

– Concentrations of salts, metals and PAHs were highest in column eluates of CDW3, particularly in U and S fractions (Figure 3). Notably, the concentrations of some salts and metals increased in U and S fractions compared to In, indicating a possible redistribution and accumulation of contaminants in the fine-grained fractions during the washing process. While concentrations in the solids of silt fractions were 2–3 times higher than in the sand fraction (see Figure 2), leaching at L/S 2 resulted in much lower concentrations in U compared to S in most cases, suggesting that metals bind stronger to finer particles (Figure 3) [23]. From the PAHs, phenanthrene, fluoranthene and pyrene showed the highest concentrations in column eluates of the different solid fractions (Figure 3). The sum of the 16 PAHs exceeded the GFS values (>0.2 µg/L) in all column eluates except those of the gravel fractions (Figure 3). The input material (In) shows concentrations that fall in between those observed for U, S and G fractions (Figure 3). The results obtained for CDW reflects the impact of material heterogeneity on contaminant leaching [24], which is highly variable in all cases.


**Table 1.** Column parameters of short and extensive leaching tests of railway ballast (RB) and three sets of construction & demolition waste (CDW) materials. In: input material, U: silt fraction, S: sand fraction and G: gravel fraction.

<sup>a</sup> Extensive leaching tests of material CDW3 only. <sup>b</sup> For the input material (In) and silt fraction (U), values represent the mixture of sample with dry quartz sand. <sup>c</sup> Total organic carbon. TOC was not measured in the gravel fraction due to expected insignificant organic carbon content. <sup>d</sup> Column parameters of the input material (In) are related to the columns performed for ions and metals analyses. <sup>e</sup> Column parameters of the silt fraction (U) are related to the columns performed for ions and metals analyses.

**Figure 3.** Concentrations of salts, metals and PAHs from short column tests at L/S 2 ratio of railway **Figure 3.** Concentrations of salts, metals and PAHs from short column tests at L/S 2 ratio of railway ballast (RB) and three sets of construction and demolition waste (CDW); WW represents concentrations in the washing water used for cleaning and separation of the input material (In) into the different size fractions silt (U), sand (S) and gravel (G). Error bars represent uncertainties in measurements.

**3 S In S In S In In**  In terms of usability of the solid waste material for different recycling purposes, the fine-solid fractions of RB and CDW are not suited for specific applications in technical constructions, which are sensitive with regard to groundwater protection (e.g., open applications with less than 1.5-m groundwater distance). These materials can be recycled only in isolated or semi-isolated applications with more than 1.5 m distance to the groundwater table and with suitable subsoil characteristics complying with the highest material classes (e.g., BM-F2 or BM-F3) [11,19]. As for the sand fraction of CDW, PAHs concentrations

] 25.1 11.5 0.60 - 9.61 5.14 1.42 - 13.4 5.59 2.62 - 13.7 4.42 1.00 - CDW3

reached limit values for the best material class RC-1 of 4 µg/L. Overall, our results proved that the least contaminated fraction is the gravel (see Figures 2 and 3), which is suitable for free re-use in all open applications in all technical constructions. Sand and silt fractions can be re-used as recycling material only in isolated or semi-isolated technical applications. Concerning contamination with PAHs, limited applications are possible only if the solid concentration limits for PAHs are met additionally, which is not always the case (PAHs exceed limits for recycling even after wet processing, e.g., CDW-2).

Solute concentrations were also measured in the washing water (WW, Figure 3) used during the separation of the solid materials into different grain-size fractions (i.e., silt, sand and gravel) at the recycling plant. The washing water showed concentrations of metals such as As, Cr, Cu and Mo up to 4.5 µg/L, 54 µg/L, 50 µg/L and 86 µg/L, respectively. The most dominant anions were Cl− and SO<sup>4</sup> <sup>2</sup><sup>−</sup> with concentrations up to 212 mg/L and 550 mg/L, respectively. Aqueous concentrations of the sum of the 16 PAHs in WW reached up to 11.7 µg/L (Figure 3), particularly in CDW samples. Overall, the washing water (WW) showed concentrations exceeding the insignificance threshold values into groundwater (GFS values, [20]; see Table 2) and the limit values (methodological background values) for salts, and some of the metals and PAHs. These concentrations are in the range of material values of higher material classes as BM-F2 or RC-3 (Table 2) [19]. Therefore, the removal of contaminants during the washing process of solid waste material is essential to ensure adequate recycling fractions.

**Table 2.** "Insignificance threshold" concentrations (GFS) and material values of the examined organic and inorganic substances. GFS values are used to identify relevant substances in principle with regard to groundwater protection [20]. Material values are used for the classification of RB and CDW into material classes, which are linked with permissible applications in technical constructions, regulated in the upcoming German recycling degree [19].


a Insignificance threshold values for groundwater protection (GFS values) [20]. <sup>b</sup> Material values with regard to technical constructions of soil materials BM-F0\* [19]. <sup>c</sup> Material values with regard to technical constructions of soil materials BM-F1 [19]. <sup>d</sup> Material values with regard to technical constructions of soil materials BM-F2 [19]. <sup>e</sup> Material values with regard to technical constructions of RC-1 defined as the highest quality construction and demolition waste [19]. <sup>f</sup> 15 PAHs, excluding naphthalene and methylnaphthalene. <sup>g</sup> Sum of PCBs (PCB-28, -52, -101, -138, -153, -180) and PCB-118. <sup>h</sup> Limit concentrations of petroleum hydrocarbons ranging from C10 to C40.

#### *3.3. Comparison between Short and Extensive Column Tests: The Importance of Compliance Testing*

Of the three sets of CDW (i.e., CDW1-CDW3), CDW3 material was selected to further examine the long-term leaching behavior of potential contaminants, as it proved to be the most contaminated solid material in L/S 2 eluates, particularly for the silt and sand fractions (Figure 3). The gravel fraction was not further examined as eluate concentrations in L/S 2 were lower than GFS values (see Figure 3 and Table 2). Figure 4 compares cumulative leaching in long-term to short-term tests at L/S 2 ratios. Figure 5 shows the grouping of salts, metals and PAHs in normalized leaching plots, and Figures 6–8 show the dynamics of the long-term leaching behavior in log-log plots.

**Figure 4.** Comparison of short and extensive leaching test results for sample CDW3 (L/S 2); solid line represents the linear regression of the data (R <sup>2</sup> = 0.92, slope = 1.33).

− − −

("salts") based on nor-

Normalized concentrations of selected groups of compounds and elements ("salts" left , "metals" middle **Figure 5.** Normalized concentrations of selected groups of compounds and elements ("salts" left panel (**a**,**d**,**g**), "metals" middle panel (**b**,**e**,**h**), and PAHs right panel (**c**,**f**,**i**) vs. liquid-to-solid (L/S) ratio). Colored lines and symbols represent observations from extensive column test of different solid fractions of CDW 3: input material (In), silt (U), and sand (S); dashed lines represent the fitted results from the advection-dispersion model (with distribution coefficients *K<sup>d</sup>* ranging from 0.28–3.64 L/kg and *α/x* ratios from 0.07–0.50 for salts and metals).

Normalized concentrations of selected groups of compounds and elements ("salts" left

, "metals" middle

ehavior of selected "salts" from different grain – – – – **Figure 6.** Leaching behavior of selected "salts" from different grain-size fractions of CDW material (CDW3) until a liquid-to-solid (L/S) ratio of 10 (dotted line L/S = 2); red diamonds: input material (mixture), blue circles: silt (<0.063 mm), orange triangles: sand (0–0.2 mm). (**a**) TSS, (**b**) pH, (**c**) Eh, (**d**) DOC, (**e**) Cl −, (**f**) NO<sup>3</sup> −, (**g**) SO<sup>4</sup> <sup>2</sup>−, and (**h**) K + .

– **Figure 7.** Leaching behavior of selected metals from different grain size fractions of CDW material (CDW3) until a liquid-to-solid (L/S) ratio of 10 (dotted line L/S = 2); red diamonds: input material (mixture), blue circles: silt (<0.063 mm), orange triangles: sand (0–0.2 mm). (**a**) As, (**b**) Cr, (**c**) Cu, (**d**) Mo, (**e**) Ni, and (**f**) Se.

– **Figure 8.** Leaching behavior of selected PAHs from different grain-size fractions of CDW material (CDW3) until a liquid-to-solid (L/S) ratio of 10 (dotted line L/S = 2); red diamonds: input material (mixture), blue circles: silt (<0.063 mm), orange triangles: sand (0–0.2 mm). (**a**) Phe, (**b**) Ant, (**c**) Fth, (**d**) Py, (**e**) BaP, and (**f**) the sum of 16PAHs.

– Long-term column tests were sampled from LS 0.1 (0.5 for silt) to L/S 10 (extensive tests) and analyzed for "salts", metals and PAHs. Cumulative concentrations (*Ccum*) were calculated from the cumulative mass released up to L/S 2 divided by the total volume of water at L/S 2. As expected, very good agreement was observed between aqueous concentrations from the short and extensive column tests at L/S 2 ratios proving that onestep short column tests are sufficient for compliance testing, and thus reduce testing time (Figure 4). Short-term column percolation tests are thus suitable for continuous internal (facility control) and continuous external quality control. Some variability was observed in the sand fraction, particularly for metals, which may be due to more complex solubility behavior relative to pH and redox conditions [21,25].

#### *3.4. Typical Release Patterns of Groups of Substances and Fitting of the Advection-Dispersion Transport Model*

The different mass-release pattern of salts, metals and PAHs observed in eluates of the extensive column tests demonstrated that substances can be grouped into rapid ("salts"), intermediate ("metals") and slow leaching substances such as PAHs [3,4]. Figure 5 shows the similar leaching behavior of DOC, Cl −, NO<sup>3</sup> − and SO<sup>4</sup> <sup>2</sup><sup>−</sup> ("salts") based on normalized concentrations of the input material as well as silt and sand fractions. Metals such as Mo, Ni, Cu and Se may also be grouped, and showed a partially slower leaching than the "salts". In general, slower leaching was observed in the silt fraction, probably due to smaller grain size and higher sorption capacity. PAHs showed decreasing concentrations only in the input material, while partly stable concentrations were observed for the silt and sand fractions. These similar leaching patterns were observed for Phe, Fth, Py and most of the other 16 PAHs (Figure 5).

As suggested earlier [3], a simple parsimonious transport model may be used to describe leaching in column tests, and to obtain average *K<sup>d</sup>* and longitudinal dispersivity (*α*) values by fitting to observed data (dashed line in Figure 5). A description of the model is provided in Appendix B. It should be noted that the fitting parameters *K<sup>d</sup>* and dispersivity (here *α/x* - *α* as a function of the length of the pack column *x*) lump together

all processes that are not accounted for in the analytical solution, such as non-equilibrium sorption/desorption, non-linear sorption and/or slow desorption, which all lead to extended tailing and thus increased "dispersivity" (e.g., *α/x* > 0.12) [3]. While the approach worked reasonably well for "salts" and "metals", PAHs did not follow the model well, probably also due to artifacts in measurements (Figure 5).

The estimated (fitted) average *K<sup>d</sup>* for salts and metals from eluates of the sand fraction were 0.28 to 0.32 L/kg, respectively, while eluates from the silt fraction resulted in the same *K<sup>d</sup>* values of 3.6 L/kg. For the input material, *K<sup>d</sup>* values of 0.7 and 1 L/kg were obtained for salts and metals, respectively. These results further support the high heterogeneity in the input material with a high fraction of silt leading to slower leaching (see also Figures 6 and 7). Dispersivities fitted as *α/x* were 0.31–0.50 in the input material and thus larger than in silt (0.14–0.15) and sand (0.07–0.16) fractions, possibly also due to the pronounced heterogeneity of the input material. Overall, the results of the fitting model indicate that in most cases leaching initially occurs at or reasonably close to equilibrium, as indicated by Grathwohl and Susset (2009), in particular for the homogeneous material fractions.

#### *3.5. Dynamics of Contaminant Leaching*

Figures 6–8 illustrate the long-term dynamics of contaminant leaching in log-log plots. With the exception of TSS and pH, the "salts" (SO<sup>4</sup> <sup>2</sup>−, NO3−, NO2−, Cl<sup>−</sup> and Na<sup>+</sup> ) showed, initially, high concentrations (up to 400 mg/L, higher concentrations than in the short column tests but lower than in washing water), which decreased by 90% after L/S 2 in the sand fraction and input material (Figure 6). Silt showed the highest and most stable concentrations for PAHs and metals, which are attributed to high sorption capacity of the fine particles (Figure 7). In the sand fraction, a delayed decline in the metals was observed in some cases. The untreated input material showed mostly low and continuously decreasing concentrations, which likely results from the heterogeneity of the material largely composed by coarser fractions, which were least contaminated by PAHs and highly polluted fine materials (see Figures 2 and 3). This probably leads to a superposition of solute leaching from different material classes and the typical power-law behavior observed for the input material in Figures 6–8. Liu et al. (2021) [12] showed that heterogeneous mixtures of materials may result in very complex contaminant release characteristics in column leaching tests, especially if materials with different degrees of contamination are concerned. For example, a rapid initial decline in concentrations followed by concentration "tailing" maybe be explained by a heterogeneous material in which a small portion of less sorbing material (low *K<sup>d</sup>* , high *Cw,eq*, low retardation) is mixed with a more strongly sorbing material (high *K<sup>d</sup>* , low *Cw,eq*, high retardation) [12].

The leaching of metals typically varies from sample to sample and likely depends on several other parameters that change over time, such as pH, redox potential and ionic strength [3,15,26]. Here, initial metal concentrations were highest in the sand fraction (in contrast to the salts); Cr and Mo showed concentrations up to 286 µg/L and 269 µg/L (at L/S = 0.1 L/kg), respectively, followed by Cu (93 µg/L; Figure 7). Generally, metal concentrations decreased again by 90% at L/S 2 in the input material and sand fraction, including As, Ni and Se.

The release pattern of PAHs also varied among the different grain-size fractions of CDW3. Initial concentrations of the 16 PAHs of the input material, silt and sand fractions were >0.2 µg/L (Figure 8). While concentrations of most PAHs were quite constant in the sand and silt fractions, indicating strong sorption; the input material showed a powerlaw behavior with continuously decreasing concentrations, as already observed for some of the metals and salts (see Figures 6 and 7). The sudden drop in leachate concentrations at L/S 10 of the sand fraction is unclear and possibly reflects an artifact during the sampling procedure.

Generally, PAHs (and metals) may be associated to suspended particles or dissolved organic matter [27–30], but since turbidity and DOC remained well below 100 mg/L (see Figure 6), respectively, this would only affect strongly sorbing compounds with *K<sup>d</sup>* values

larger than 10,000 L/kg [13]. DOC was always below 50 mg/L and continuously decreased to less than 10 mg/L (see Figure 6) in all fractions, which is not reflected in the rather stable concentrations of PAHs, e.g., in sand and silt fractions. Similarly, metals such as copper, which are known to form complexes with DOC showed fairly stable concentrations in the silt fraction, while DOC decreased rapidly. High TSS values were observed for silt and the input material, while the sand fraction showed very low and declining TSS values (see Figure 6). TSS in leachate of the silt fraction was quite stable and even showed an increase at L/S 10, while TSS of the input material dropped from 100 mg/L to 10 mg/L, which in principle could have affected leaching of high molecular weight PAHs (Fth, Pyr, BaP). Since all PAHs showed a similar leaching behavior, and concentrations in the sand fraction were higher than in the input material, particle facilitated transport seems not to play a major role (maybe with the exception of BaP, which, however, has the lowest concentrations and does not significantly contribute to the sum of PAHs).

#### **4. Conclusions**

Wet processing after crushing of CDW and RB produces approximately 25 % of silt and sand, respectively, whereas the gravel fraction is usually around 50 %. Coarse-grained fractions (gravels) generally fulfilled legal standards for a free reuse in open technical applications (landscaping, etc.), while the sand fractions still showed concentrations which limit their reuse to specific technical applications. Fine-grain fractions (silt) are still contaminated and only allow limited re-use in (semi-) isolated applications, or require landfilling. This is also reflected in concentrations in solids and aqueous leachates up to L/S 2 (Figures 2 and 3).

Results from the short leaching tests showed to be comparable with the cumulative concentrations from the extensive column tests (up to L/S 2 L/kg; Figure 4). Thus, short leaching tests are suitable for compliance testing where concentrations can be compared to threshold values in order to select various material fractions for different recycling applications. Extensive column leaching tests showed, particularly for salts and some metals, a highly dynamic contaminant release with a decline to less than 10% of the initial concentration at *L/S* 2 for the sand fraction and input material. The silt fraction showed quite stable concentrations up to *L/S* 10, probably due to high sorption capacities for metals and PAHs. The leaching behavior of organic and inorganic substances from highly heterogeneous materials (i.e., "input material" of CDW 3) reflects their complex composition, making leaching patterns difficult to predict. As observed in earlier studies, a "typical" leaching behavior of highly soluble substances such as Cl<sup>−</sup> and SO<sup>4</sup> <sup>2</sup>−, and metals such as Cu and Mo allows their grouping and can fit with simple transport models. Overall, short column leaching tests provide important information for decision making on the recycling of waste material. Future similar studies may help to optimize processing of mixed solid waste for higher recoveries of material fractions suitable for recycling.

**Supplementary Materials:** The following supporting information can be downloaded at: https:// www.mdpi.com/article/10.3390/ma15030858/s1, Table S1: Chemical analyses of solids and leachates.

**Author Contributions:** Conceptualization, M.P.-E., B.S. and P.G.; validation, B.S. and P.G.; investigation, M.P.-E., B.S. and P.G.; writing—original draft preparation, M.P.-E.; writing—review and editing, M.P.-E., B.S. and P.G.; visualization, M.P.-E.; supervision, B.S. and P.G. All authors have read and agreed to the published version of the manuscript.

**Funding:** The first author acknowledges the support by the Mexican National Council for Science and Technology (CONACyT). We acknowledge support by Open Access Publishing Fund of University of Tübingen.

**Institutional Review Board Statement:** Not applicable.

**Informed Consent Statement:** Not applicable.

**Data Availability Statement:** The data presented in this study are available in this article as Supplementary Information.

**Acknowledgments:** M.P.E. acknowledges Thomas Wendel for his support during the performance of the column tests. We thank Renate Seelig, Sara Cafisso, Larissa Lohmüller, Annegret Walz and Bernice Nisch for their laboratory assistance.

**Conflicts of Interest:** The authors declare no conflict of interest.

#### **Appendix A. Concentration Measurements**

**Table A1.** Solid concentrations of railway ballast (RB) and three sets of construction and demolition waste (CDW) from different solid fractions. In: input material, U: silt fraction, S: sand fraction and G: gravel fraction. Values exceeding precautionary values in bold.


<sup>a</sup> Petroleum hydrocarbons of chain C10-C40. <sup>b</sup> Sum of PCBs (PCB28, PCB52, PCB101, PCB138, PCB153, PCB180). Precautionary values are given for PCB6 and PCB118 [19]. <sup>c</sup> Threshold concentration based on precautionary values in soils (BM-0 Sand) [19]. <sup>d</sup> Precautionary value of petroleum hydrocarbons (PHCs of C10-C40) based on limit value of "material class" BM-0\* [19].

**Table A2.** Measured concentrations of herbicides in short leaching test of railway ballast (RB) on different solid fractions. In: input material, U: silt fraction, S: sand fraction and G: gravel fraction.


a Insignificant threshold values into groundwater [20]. <sup>b</sup> Threshold concentration for recycling railway ballast material (GS-0) [19].

#### **Appendix B. Long-Term Leaching Behavior Described by the Advection-Dispersion Equation**

Column leaching tests represent the percolation of water through different types of solid materials. Initially, the column is saturated so that equilibrium conditions can be achieved rather rapidly (<5 h) [13]. The drop in concentrations is given by a change in non-equilibrium conditions leading to an extended tailing of low concentrations. The advection-dispersion model allows for the description of the movement of the front of clean water through the column as:

$$\frac{\partial \mathcal{C}}{\partial t} = \frac{D}{R} \frac{\partial^2 \mathcal{C}}{\partial \mathbf{x}^2} - \frac{v}{R} \frac{\partial \mathcal{C}}{\partial \mathbf{x}} \tag{A1}$$

where *C* is the solute concentration, *D* is the longitudinal dispersion coefficient [m<sup>2</sup> s −1 ], *v* is the average flow velocity [m s−<sup>1</sup> ], *t* is time, *x* is the length of the column [m] and *R* denotes the retardation factor [-], defined as:

$$R = 1 + K\_d \frac{\rho}{n} \tag{A2}$$

where *K<sup>d</sup>* denotes the distribution coefficient [L/kg], defined as the ratio of concentrations in the solids to the aqueous concentrations (*Cs*/*Cw*), *ρ* is the dry bulk density [kg/L] and *n* is the porosity [-]. The advection–dispersion model assumes local equilibrium conditions, but previous studies demonstrated a reasonably well fit with the early leaching behavior [3,4]. This model is solved using the analytical solution for the movement of the front of clean water through the column [31] and expressed based on the dynamic liquid to solid ratio (L/S) as:

$$\frac{\mathcal{C}}{\mathcal{C}\_{0}} = 1 - 0.5 \left[ \text{erfc} \left( \frac{\mathcal{K}\_{d} - \mathcal{L}S}{2 \sqrt{\frac{a}{\mu} \left( \frac{\mathcal{U}}{\rho} + \mathcal{K}\_{d} \right) \mathcal{L}S}} \right) + \exp \left( \frac{\mathbf{x} \left( 1 - \frac{1}{\mathcal{R}} \right)}{a} \right) \text{erfc} \left( \frac{\mathcal{K}\_{d} - \mathcal{L}s}{2 \sqrt{\frac{a}{\mu} \left( \frac{\mathcal{U}}{\rho} + \mathcal{K}\_{d} \right) \mathcal{L}S}} \right) \left( 1 - \frac{\mathcal{C}\_{\text{min}}}{\mathcal{C}\_{0}} \right) \right] \tag{A3}$$

where *Cmin* is the minimum concentrations usually detected at L/S 10 L/kg, and LS is the amount of water percolated through the column after *t* time relative to the dry weights of the solids in the column (= *v n t/x ρ*). The last term in brackets has been here added to fit the late data of the leaching tests, which show substantial tailing. Eq. B3 accounts for the initial displacement of low sorbing (high soluble) compounds during first flooding of the column. The model is fitted to measured data using MATLAB (v R2021b) and the function *lsqcurvefit*. From the fit, retardation factors (*R*) and distribution coefficients (*K<sup>d</sup>* ) were calculated. In addition, the longitudinal dispersivity was fitted as a function of *x* (i.e., *α/x*). Maximum values of *K<sup>d</sup>* and *α* were set to 100 L/kg and 1 m, respectively.

#### **References**


### *Article* **Environmental Impact of Geosynthetics in Coastal Protection**

**Philipp Scholz 1 , Ieva Putna-Nimane 2 , Ieva Barda 2 , Ineta Liepina-Leimane 2 , Evita Strode 2 , Alexandr Kileso 3,4 , Elena Esiukova 3 , Boris Chubarenko 3 , Ingrida Purina <sup>2</sup> and Franz-Georg Simon 1, \***


**Abstract:** Geosynthetic materials are applied in measures for coastal protection. Weathering or any damage of constructions, as shown by a field study in Kaliningrad Oblast (Russia), could lead to the littering of the beach or the sea (marine littering) and the discharge of possibly harmful additives into the marine environment. The ageing behavior of a widely used geotextile made of polypropylene was studied by artificial accelerated ageing in water-filled autoclaves at temperatures of 30 to 80 ◦C and pressures of 10 to 50 bar. Tensile strength tests were used to evaluate the progress of ageing, concluding that temperature rather than pressure was the main factor influencing the ageing of geotextiles. Using a modified Arrhenius equation, it was possible to calculate the half-life for the loss of 50% of the strain, which corresponds to approximately 330 years. Dynamic surface leaching and ecotoxicological tests were performed to determine the possible release of contaminants. No harmful effects on the test organisms were observed.

**Keywords:** geosynthetics; geotextiles; dynamic surface leaching test; artificial ageing; marine littering

#### **1. Introduction**

Geosynthetics are widely used in coastal protection. Their application areas are soil reinforcement, the stabilization of ballast layers, filtration, the waterproofing of dams and canals, and scour protection (e.g., for piles of offshore wind energy plants). The application of geosynthetics in coastal protection has huge economic benefits, such as savings via substitutions of or reductions in selected soil materials, ease of installation, increased speed of construction, life cycle cost savings through improved performance (by increased longevity or reduction in maintenance), and improved sustainability in terms of conserving natural environments as compared to alternative designs [1,2]. It is commonly accepted that geosynthetics which are adequately stabilized with antioxidants (e.g., sterically hindered amines) will last in underwater constructions with limited oxygen supply and temperatures at constantly low levels for at least 100 years.

However, after the end of service lifetime, geosynthetics could be a source of plastic debris in aquatic systems if the construction which the geosynthetic is a part of is not dismantled. Further, additives which are needed as plasticizers or antioxidants could be emitted, with detrimental influence on the environment [3]. The loss of additives is intimately related to the aging of the geosynthetic products. These are the reasons that public authorities are concerned about the approvability of engineering projects using geosynthetics in aquatic systems.

The long-term stability of geotextiles is usually investigated with relation to mechanical stability, which must fulfill certain requirements after aging. Various methodologies are available (e.g., elevated temperatures or increase in oxygen pressure) to accelerate aging in the laboratory [4]. Mechanical properties, such as tensile strength, investigation of chemical

**Citation:** Scholz, P.; Putna-Nimane, I.; Barda, I.; Liepina-Leimane, I.; Strode, E.; Kileso, A.; Esiukova, E.; Chubarenko, B.; Purina, I.; Simon, F.-G. Environmental Impact of Geosynthetics in Coastal Protection. *Materials* **2021**, *14*, 634. https:// doi.org/10.3390/ma14030634

Academic Editor: Qing-feng Liu Received: 4 January 2021 Accepted: 25 January 2021 Published: 29 January 2021

**Publisher's Note:** MDPI stays neutral with regard to jurisdictional claims in published maps and institutional affiliations.

**Copyright:** © 2021 by the authors. Licensee MDPI, Basel, Switzerland. This article is an open access article distributed under the terms and conditions of the Creative Commons Attribution (CC BY) license (https:// creativecommons.org/licenses/by/ 4.0/).

oxidation reactions by infrared spectroscopy, and the residual content of stabilizers are typical parameters tested on aged samples [5]. The investigation of the possible environmental impact of the application of geosynthetics in aquatic systems is therefore hardly possible with virgin polymer material. Consequently, polymers must be artificially aged, which is best accomplished with environmental simulation chambers enabling accelerated ageing. In the case of geosynthetics in hydraulic engineering besides oxidation, mechanical stress (e.g., by tidal and wave action, abrasion by sand) and microbiological interactions (the formation of biofilms, etc.) [6] play significant roles and must be considered.

There are only a few investigations on the degradation behavior of geotextiles in marine environments [7,8]. According to these, exposure to UV light has a higher impact on the material properties in comparison to seawater immersion and tidal action. The importance of the stabilization of the polymers was strengthened. It can be expected that the degradation processes of geotextiles are similar to the processes of other plastics reaching the marine environment because they are made from the same types of polymers. Plastic waste exposed to environmental conditions begins to degrade slowly under the impact of temperature and UV radiation [9], generating a large number of macro-, micro, and nano-particles. These particles are freely transported by water flows and have adverse effects on the environment [10,11]. One of the key factors which determines the fate of microplastics in the environment is the density of polymers. The specific density of microplastic can vary significantly depending on the polymer type, technological processes of its production, additives, weathering, and biofouling [12,13]. With time, most floating plastics become negatively buoyant due to both biofouling and the adherence of denser particles and sink to the sea floor [13,14]. Thus, the seabed becomes the ultimate repository for microplastic particles and fibers [15,16]. The evaluation of the contamination level is complicated, not only because of the difficulty of the sampling of sea bottom sediment, but also due to the difficulty of the extraction of small plastic particles from marine deposits.

The project Environmental Impact of Geosynthetics in Aquatic Systems (EI-GEO) [17] aims at the investigation of whether geosynthetics in hydraulic engineering applications could be a source of microplastic or other contaminants in the aquatic environment. Whereas the behavior of geosynthetics in landfill engineering has been well studied and documented for decades [18], little is known regarding applications such as coastal protection or scour protection for off-shore wind energy plants. However, due to the rapid expansion of offshore wind energy, rising water levels, and more extreme weather conditions as a result of climate change, more and more hydraulic engineering projects will be realized in the future.

Construction with geosynthetics boasts various advantages, but it has to be ensured that there is no negative environmental impact from the application of geosynthetics in hydraulic engineering. It is expected that any effect will be visible only in the long term because the virgin raw material used for the production of geosynthetics has almost no release of particles or substances relevant to the environment [19].

Partly from improper material selection and partly from non-professional handling, debris from geosynthetic material can be found on the shore today. Therefore, a field study with sampling and monitoring was performed and the magnitude of this pollution was evaluated (objective 1). Further, an accelerated ageing method was performed to derive the requirements for geosynthetics in hydraulic engineering. The testing of mechanical properties was performed with virgin and artificially aged geosynthetics (objective 2). Finally, leachates of artificially aged geosynthetics were used in ecotoxicological tests, which are essential tools to evaluate the environmental impacts of the pollutants released by geosynthetics during ageing (objective 3).

#### **2. Materials and Methods**

The applications of geosynthetics in hydraulic and coastal engineering such as revetments, dyke constructions, or geotextile containers for scour prevention are described in detail elsewhere [1]. For the present study, a multifunctional geotextile for separation, filtration, and protection made of white polypropylene was selected as a test material for the investigations. The mass per unit area was 600 g m−<sup>2</sup> , the thickness was 5 mm, and the water permeability was 3 <sup>×</sup> <sup>10</sup>−<sup>2</sup> m s−<sup>1</sup> . The material, produced in Germany, is commercially available and widely used for geomembrane protection or for the production of sand container bags.

#### *2.1. Accelerated Ageing Using Autoclave Test*

Autoclave tests following DIN EN ISO 13438:2005 (method C) [19] were performed under a pure oxygen atmosphere with pressures between 10 and 50 bar, at temperatures between 30 and 80 ◦C, and with durations in the range of 14 to 143 days. An overview on the performed ageing experiment is given in Table 1. It is important to notice that the test specimens were completely immersed in tap water and the exposure of autoclaves was carried out based on the time-dependent degradation of the mechanical properties of the polypropylene geotextiles. Five PP specimens (250 <sup>×</sup> 50 mm<sup>2</sup> ) were placed in the autoclaves in tap water. The use of artificial seawater was not possible due to the risk of chlorine-induced corrosion at high oxygen pressures. In order to reach thermal equilibrium, the autoclaves were left for 48 h in electronically controlled heating systems before the start of the tests. Hence, single specimens were removed in succession after different ageing periods. Then, the tensile strength was determined accordingly. Two measurements were carried out for each duration of aging. All the tensile test measurements were performed with a Zwick tensile testing machine (Zwick-Roell, Ulm, Germany) (ZPM Model 1464 with testXpert II software (Version 3.31, Zwick, Ulm, Germany)) with a 5 kN force sensor. The tensile tests were performed in an air-conditioned environment at 23 ◦C and a relative humidity of 50%. For the tensile test measurements, a clamping length of 50 mm and a test speed of 50 mm/min were chosen. Each sample was attached to a sandpaper to avoid sliding during the tensile test.


**Table 1.** Duration of accelerated ageing in autoclaves in days at 5 different temperatures and pressures.

Figure 1 shows a sketch of the autoclave test equipment along with all the instruments and monitoring devices used. The temperature and the pressure were observed and recorded every 15 min using an electronic data recorder (Eurotherm 6100) (Eurotherm, Limburg, Germany). The temperature of the autoclave was controlled by an external heating jacket with a separate PT100 temperature sensor connected to a PID temperature controller (Eurotherm 2216E) (Eurotherm). The heating power line was equipped with an electrical contact controlled by the internal temperature monitoring to prevent overheating of the system. The safe and reliable operation of the autoclaves requires the control and monitoring of the relevant parameters, especially for long-term experiments. All the relevant instruments and transducers were calibrated in order to obtain reliable and reproducible results.

**Figure 1.** Schematic view of autoclave test equipment (**left**), closing the cover plate of the autoclaves (test rig with two autoclaves).

#### *2.2. Dynamic Surface Leaching Test*

Dynamic surface leaching tests (DSLT) were performed on the geosynthetic materials according to the CEN/TS 16637-2 leaching method [20]. The DSLT corresponds to a tank test for the assessment of the surface-dependent release of dangerous substances and is suitable for monolithic construction products. The test specimens were eluted using demineralized water at a defined water/surface ratio (L/A) and a water exchange at several fixed time intervals (6 h, 1 d, 36 d). The L/A ratio was set to 80 L/m<sup>2</sup> in CEN TS 16637-2, but can be reduced to 25 L/m<sup>2</sup> for plate-like products. Tests were performed at 23 ± 2 ◦C, room humidity 50 ± 5%, in the darkness. Two plates were eluted per coating system to obtain enough eluate volume for all the ecotoxicological tests. Each plate was individually placed in a tank and the eluates of the same fraction were combined and well mixed before aliquoting them for ecotoxicological analysis.

#### *2.3. Ecotoxicological Testing*

≤ ≥ Internationally agreed and accepted ecotoxicity test methods have been performed to demonstrate the impact of chemicals and other pollutants on the environment and determine the potential damage to organisms and the function of ecosystems [21–23]. Ecotoxicity tests consisted of two acute and one chronic test with organisms from different levels of aquatic food chains. The ecotoxicity test conditions, growth media, dilutions, and replication are summarized in Table 2. The test eligibility criteria for the *Daphnia magna* test is ≤10% immobile organisms in the control treatment and an ≥80% survival for the *Hyalella Azteca* test. For the *Desmodesmus subspicatus* test, control batch absorption measurements should indicate the exponential growth of algal cells, the variation coefficient (CV) of the growth rate in the control replicates should not exceed 5%, and the pH in the control should not increase during the test by more than 1.5 relative to the pH of the growth medium.


**Table 2.** Ecotoxicity test conditions summary.

\* ADaM: Aachener Daphnia Medium.

#### *2.4. Continuous Visual Scanning (Field Study)*

Since the fragments of plastics and geosynthetic materials were unevenly distributed on the beach, the use of a selective area technique for their search—such as, for example, for anthropogenic debris [24] and microplastics [25]—will not yield results. To analyze the pollution of the beaches at the Southeastern Baltic within the Kaliningrad Oblast (Russia), a continuous visual scanning technique [26] was applied which assumes a continuous passage of a group of several people along the entire coastline, covering the entire width of the beach from the shoreline to the foredune (or cliff).

The width of the beaches of the Kaliningrad Oblast ranges from almost 0 to 188 m and the average value is 30 m, so the group usually included three people. The beach was divided into three control strips, each member of the group controls his strip and even tries to capture the edge of the neighboring zone for a complete scan of the entire beach. During the day, the group could walk 7–10 km, and such monitoring surveys were carried out in 2018.

Each detected plastic or geosynthetic fragment with a size larger than 3–5 cm was attributed to the different type of origin (see Results section), dimension scale (length and area), number of the coastal subsection where this sample was found, and position on the beach (in % of the beach width). Next, photographs were taken and, if necessary, the sample was saved for further laboratory analysis.

#### **3. Results**

#### *3.1. Field Study on Kaliningrad Oblast Shore (Russia)*

During the surveys of the beaches of the Kaliningrad Oblast (Figure 2) in 2018, a large amount of remnants of geosynthetic materials that are used in coastal protection structures [27] were found. In addition, there was extensive contamination from other building support materials—e.g., geotextile FIBC (Flexible Intermediate Bulk Container) bags ("big bags") and the remains of fishing nets, ropes, and car tires.

**2021**, , x FOR PEER REVIEW 6 of 13

**Figure 2.** Shoreline of the Kaliningrad Oblast (Russia) in the Baltic Sea including the Sambian Peninsula (quadrangle). Source: OpenStreetMap.

In 2018, 3485 samples were collected from the beaches which, by origin, belonged to several types of materials: geotextiles, geocells, geogrids, plastic coating from gabions, and geotextile big bags. The integral amount of remnants of geotextile objects was more than 190 m<sup>2</sup> and the integral length of the geotextile braids from gabions coating was about 100 m [28].

The occurrence of geosynthetic remnants varies greatly along the entire shore of the Kaliningrad Oblast. The northern shore of the Sambia Peninsula accounts for 66% of the remains found, 31% for the beaches of the Curonian Spit National Park, and only 3% was found on the beaches of the western shore of the Sambia Peninsula and the Vistula Spit. Among the remains of geosynthetic materials found, the largest number was braid from gabions (44%) and geocontainers (43%), pieces of geotextile accounted for only 12%, and the remaining 1% was made up of remnants of geocells and geogrids.

The performed primary statistical analysis on the occurrence of the number of pieces per 1 kilometer for various morphodynamic segments of the coast of the Kaliningrad Oblast (Vistula Spit, western and northern shores of the Sambia Peninsula, Curonian Spit) showed that the main pollution occurs on the northern shore (see Table 3). Considering the average size of one piece of geotextile (0.9 m<sup>2</sup> ), gabion coating (7.4 cm), big bag (0.3 m<sup>2</sup> ), and geocell (0.06 m<sup>2</sup> ), it is obvious that the remnants of geotextile and "big bags" were the mostly visible litter on the beach.

**Figure 3.** Photographs of samples collected during the field study: (**a**) + (**b**): aged plastic coating of wires in gabions' (**c**) debris from geocell; (**d**): debris from big bag.



Note: Numbers are given in pieces/km, while pieces have very different linear sizes (see Figure 3 for examples).

This fact that the northern shore of the Sambian Peninsula is mostly littered correlates well with the location of engineering structures using geosynthetic materials, most of which are located on the northern shore of the Sambian Peninsula [27]. In addition, the main accumulation of residues of geosynthetic materials is observed in the areas adjacent to these engineering structures. On the Curonian Spit (north from the Sambian Peninsula), a large amount of geosynthetic remnants was also found, which were probably brought here by alongshore currents [29]. The occurrence of residues on the Vistula Spit (south from the Sambian Peninsula) and on the western coast of the Sambia Peninsula is low due to the current structure in the eastern part of the Gulf of Gdansk [30].

ε Gabion coating was found quite often (see Table 3). This came from the plastic coating of the wire used for the gabion's support structure. Obviously, this coating is not weatherproof. A support structure made of stainless steel or Zn-plated wires would not need a plastic coating but is, however, more expensive. Geotextile remnants came from

partly destroyed coastal protection structures which stay without proper maintenance during long time. Geocells were found rarely, they were from several locations, where storm events destroyed lawn on the slopes of foredune wall prepared using geocells. Debris from big bags was found often as well. However, these woven geotextiles are rather used for transport of building materials or short-term applications than for coastal protection systems. Occurrence can therefore be attributed to improper waste management. **2021**, , x FOR PEER REVIEW 8 of 13

#### *3.2. Tensile Tests after Accelerated Ageing Using Autoclave Test*

The elongation and force of break of the test specimens were measured on a tensile testing machine. The retained elongation Rε at break is measured as a function of time (and temperature and oxygen pressure) and is expected to be influenced by the ductile–brittle change which is a service lifetime criterion for the geotextile. The retained elongation Rε is defined as follows: ε <sup>ε</sup> ε ε

$$\mathbf{R}\_{\varepsilon} = 100\% \text{ o } \varepsilon\_{\text{e}} / \varepsilon\_{\text{c}} \tag{1}$$

with ε<sup>e</sup> then initial elongation at break and ε<sup>c</sup> the elongation of the exposed specimen.

The results are displayed in Figure 4. It is clearly visible that increasing temperature leads to a more pronounced decay of the mechanical properties. The loss of retained elongation proceeds with the duration of the exposure, which is visualized in Figure 4 by different gray scales of the respective symbols (bright to dark). The influence of pressure is lower. Experiments performed at 40 and 50 bar show higher values for retained elongation because the temperature was 30 ◦C and 40 ◦C, respectively.

ε **Figure 4.** Retained elongation Rε measured after exposure in autoclaves as a function of temperature (**left**) and pressure (**right**). The duration of exposure is visualized by the gray scale of the symbols. Note that at 80 ◦C experiments at three different pressures (10, 20, and 30 bar, different symbols) were performed.

− The aging of polymers is caused by oxidation. The thermo-oxidation of PP can be defined as an in-chain radical mechanism. The latter generates hydroperoxides more rapidly than they decompose, which strengthen its strong auto-accelerating character. A detailed description of the oxidative aging of polymers is given by Verdu [31]. The accelerated ageing in the autoclaves is a function of temperature and pressure with a (pseudo-)first-order rate constant k (s−<sup>1</sup> ). The temperature and pressure dependence of the oxidation reaction can be approximated by an modified Arrhenius equation (consideration of pressure dependence) [32,33]:

$$
\ln \frac{\varepsilon\_{\text{e}}}{\varepsilon\_{\text{c}}} \sim \ln \frac{\mathbf{c}\_{0}}{\mathbf{c}} = \text{A} \cdot \exp \left( \frac{-\mathbf{E}\_{\text{a}} + \mathbf{C} \cdot \mathbf{p}}{\mathbf{R} \, \text{T}} \right) = \text{jk} \,\left( \text{T}, \, \mathbf{p} \right) \,\text{t}, \tag{2}
$$

−

− − − − with frequency factor A (s−<sup>1</sup> ), activation energy E<sup>a</sup> (J mol−<sup>1</sup> ), pressure factor C (J mol−<sup>1</sup> bar−<sup>1</sup> ), universal gas constant R, and temperature T (K).

<sup>−</sup> τ

−

The term ln c0/c is usually related to the fate of a substance in a chemical reaction. Here, it is approximated by the loss of mechanical properties and describes the progress of the oxidation and thus degradation of the material without knowing exact concentration of oxidized and non-oxidized polymer material. The experimental data displayed in Figure 4 were fitted with the Solver module in Microsoft Excel (solver method GRG non-linear) (Office 365 for Enterprise). Starting values for activation energy E<sup>a</sup> (80,000 J mol −1 ) and frequency factor A (6 × 10 8 s −1 ) were taken from the literature [34]. As a result, k (T, p) was fitted to 0.5 s <sup>−</sup><sup>1</sup> at T = 298 K and pO2 = 0.21 bar. The half-life τ at 298 K and 0.21 bar oxygen pressure, i.e., the time were 50% of the mechanical properties are lost under ambient conditions, can be calculated from ln2/k.

$$
\pi = \ln 2 / \text{k} = \text{330 years} \tag{3}
$$

This result is in the same order of magnitude as the results from Hausmann et al. for woven polypropylene geotextiles [34] (483–795 years). Fitted pressure factor C was 146 J mol <sup>−</sup><sup>1</sup> bar −1 , so the activation energy E<sup>a</sup> in the exponential tern in Equation (2) is reduced by 7300 J mol −1 (<10%) at 50 bar oxygen pressure in the autoclave experiment. As stated above, temperature has the strongest influence on the accelerated ageing in the autoclaves, even at highest possible pressure of 50 bar. However, it must be mentioned at this point that the samples are immersed in tap water so that the samples are exposed to the dissolved oxygen in water which is proportional to the partial pressure of oxygen above the liquid (Henry's law). Henry's law solubility constant is substance specific and a function of temperature. An equation to calculate the concentration of dissolved oxygen caq in water between 273 and 616 K and pressures up to 60 bar was presented by Tromans [35] and reviewed by Sander [36]. For 50 bar and 353 K, the caq is 3.97 × 10 <sup>−</sup><sup>2</sup> mol kg −1 .

#### *3.3. Ecotoxicity Tests*

To evaluate the geosynthetic leachate ecotoxicity, a combination of bioassays was applied—both acute and chronic tests and organisms representing two trophic levels were used. Such an approach has advantages over individual component analysis and testing because it can disclose mixture effects.

The algae growth inhibition test was conducted at five volume/volume percent concentrations—5.9%, 11.8%, 23.6%, 47.2%, and 94.3%. Inhibition is evaluated by the reduction in specific growth rate relative to the cultures of the control. Samples Fraction 1 + 2 and Fraction 7 after 72 h exposures did not indicate algae growth inhibition even at the highest test concentration (Figure 5).

**Figure 5.** Algae growth response after 72 h (optical density measurements at 680 nm, **left**), *Hyalella. azteca* survival after 14 days (**right**). (NA: not analyzed, right).

The results of an acute *Daphnia magna* test showed the toxicity of Fraction 1 + 2 only at 100% concentration, causing 7.1% daphnia mortality after 24 h and 54% of cladocera mortality after 48 h exposure (Figure 6). However, there was no toxic effect observed when ADaM media microelements were added to the highest concentration. Fraction 7 did not cause any effects on *D. magna* survival during the test. **2021**, , x FOR PEER REVIEW 10 of 13

**Figure 6.** Survival of *Daphnia magna* after 24 h and 48 h.

Although the acute ecotoxicity test results of amphipod *Hyalella azteca* showed the higher toxicity of Fraction 1 + 2 than Fraction 7, no significant differences in toxicity between both samples after 14 days exposure were detected (Figure 5). In the 100% concentrate samples, an 88% mortality of amphipods was detected in Fraction 1 + 2 after 48 h exposure, while toxicity of Sample 7 increased only after one-week exposure. LC<sup>50</sup> for Fraction 1 + 2 was 83%, while Fraction 7—LC<sup>50</sup> was at 89%.

Measurements of pH showed an increase by 0.5 units after the 14-day test period, while the oxygen concentration stayed uniform more than 8.00 mg/L all test period. Ammonium concentration during the test did not reach higher than 20 mg/L (ISO 16303:2013 standard mentioned 96 h LC50 ammonium could be 20 mg/L to >200 mg/L [21]).

#### **4. Discussion**

The loss of additives, such as plasticizers and antioxidants, during the ageing of geotextiles potentially can add to the concentrations of hazardous substances in the water. This is discussed in a study from South Korea, where more than 200 different chemicals were identified in plastic marine debris and respective new products [37]. Another consideration is that base structure forming polymers gradually degrades to microplastic particles, and as such can be ingested by heterotrophs or interfere with algal photosynthesis [3]. However, ecotoxicological test results in this research did not show significant toxicity of geotextile leachates to water organisms. In case of microalgae, the test samples showed even nutritive properties, as an increase in microalgae concentration was observed during the 72 h of the test. Currently, there is limited research in the field of geosynthetic ecotoxicity, but a study evaluating the environmental safety of construction products also found that geosynthetic PET multifilament yarns and polyamide monofilament with PP fleece coating, have low toxicity [38]. Results indicate that the algae species *Desmodesmus subspicatus* that were also used in our study are slightly less sensitive than the algae *Raphidocelis subcapitata* and daphnia [39].

A concentrated sample of Fraction 1 + 2 (100%) caused mortality of *Daphnia magna.* However, if test sample was spiked with minerals from ADaM growth media, no mortality was observed. No mortality was observed in other test sample dilutions, neither in Fraction 1 + 2, nor Fraction 7. The results indicate that deionized water used in DSL tests might bias the ecotoxicity tests by adding hypoosmotic stress to low toxicity of test media. Concentrated samples (100%) of Fraction 7 did not caused mortality of organisms. These results suggest that the toxicity of additives is decreasing with time and dilution, also indicating that osmotic stress alone does not cause mortality [40].

A lethal concentration (LC50) was calculated only for amphipods *Hyallela Azteca*. However, the LC<sup>50</sup> at 83% and 89% concentrations can be considered as very low toxicity [41]. As geotextiles in hydraulic engineering are exposed to intensive water exchange, no toxic effects in the environment will be observed. However, even though within the tests with *Daphnia magna*, *Hyalella aztecal* and *Desmodesmus subspicatus* negative effects were not detected, the risk that long-term harsh climate conditions pose an impact on the release and migration of particles as well as hazardous substances cannot be excluded completely (referring to objective 3).

Service lifetime of geotextiles with state-of-the-art stabilization is far above 100 years, which was shown in the present study with accelerated ageing at elevated temperatures and oxygen pressures. The improper installation of the geotextiles and the lack of service and maintenance after extreme weather events could cause the failure of the engineered structures and, as a result, the pollution of the environment by remnants of geosynthetic materials [42]. The successful application of geotextiles in coastal protection depends on the selection of a suitable material and proper installation and maintenance (referring objective 2).

The field study performed at the shore of Kaliningrad Oblast (Russia) demonstrated that debris from plastic and geotextile materials is found in the environment [27,42]. The remnants of the geosynthetic materials are found not only at the beaches of the Kaliningrad Oblast, but at the neighboring beaches of Lithuania [43]. Some of the found objects could be attributed to unsuitable material selection (gabion coating) or improper waste management. Considering that any damage, even partial, of the coastal protective constructions using geosynthetic material could lead to the littering of the beach or the sea, specific attention is needed for the maintenance of such constructions (referring objective 1).

**Author Contributions:** Conceptualization, F.-G.S. and B.C.; Writing Original Draft, F.-G.S.; Data Acquisition, P.S., A.K., I.P.-N., E.E., I.B., I.L.-L. and E.S.; Visualization, Data Interpretation, Writing— Review & Editing P.S., B.C., I.P.-N. and I.P.; methodology, F.-G.S., B.C. and I.P.; Funding Acquisition, F.-G.S., B.C. and I.P. All authors have read and agreed to the published version of the manuscript.

**Funding:** This research was funded within the ERANET-RUS plus joint project EI-GEO, ID 212 (RFBR 18-55-76002 ERA\_a, BMBF 01DJ18005, ES RTD/2018/21 VIAA Latvia) and supported by ERDF 1.1.1.2. post-doctoral project No. 1.1.1.2/VIAA/3/19/465 and theme 0149-2019-0013 of the Shirshov Institute of Oceanology (instrumental support of the field study).

**Institutional Review Board Statement:** Not applicable.

**Informed Consent Statement:** Not applicable.

**Data Availability Statement:** Data sharing is not applicable to this article.

**Conflicts of Interest:** The authors declare no conflict of interest.

#### **References**

