Contaminants, mutagenicity and toxicity in the surface waters of Kyiv, Ukraine

Abstract

Kyiv is Ukraine’s capital and largest city. Home to 3 million people, this area has a rich history of agriculture and industry. The Dnieper River is Ukraine’s largest river and it passes through the center of Kyiv. Little information on emerging and legacy compounds or their toxicity exists for this area. Water and biota were sampled for PAHs, PCBs, metals and emerging contaminants including pharmaceuticals, personal care products and newer synthetic pesticides. The effects of surface waters in the Dnieper were evaluated using the Ames, chronic and acute daphnia and a ciliate (Colpoda stennii) assay. Elevated concentrations of legacy contaminants were found in fish samples indicating that these compounds are bioavailable. Concentrations of legacy and emerging contaminants were found in seven stations near the municipal water treatment plant (MWTP) and receiving waters of the Dnieper River. The MWTP appeared to remove some of the emerging contaminants, however the legacy compounds (PCBs and PAHs) were not affected by the MWTP and appeared to be more wide spread indicating a number of sources to the Dnieper River. Acute and chronic toxicity were associated with the influent and effluent of the MWRP, however mutagenicity was noted in surface waters throughout the Dnieper River including upstream of the MWTP. This study provides the first snapshot of possible human health and ecological risks associated with surface waters of the Dnieper. More research on seasonal changes and sources of toxicity, mutagenicity and contaminants would aid in completing a more comprehensive risk assessment of surface waters of the Dnieper River.

Introduction

Ukraine is a country of 44 million people with modern industrial cities as well as productive, fertile agricultural areas. Ukraine’s capital and largest city, houses close to three million people, is the eighth most populous city in Europe (http://www.citymayors.com/features/euro_cities1.html), and has a strategic position between eastern and western Eurasia. Historically, Ukraine served as a center for agriculture and industry during its time as part of the Union of Soviet Socialist Republics (USSR or Soviet Union) leaving a legacy of contaminants (e.g., DDT, PCBs, metals and PAHs) in freshwater and marine systems (Burgess et al., 2011; Burgess et al., 2009; Vystavna et al., 2012). Kyiv, Ukraine’s largest river, the Dnieper, runs thought the center of Kyiv. The Dnieper may have undetected emerging contaminants entering its surface waters via treated effluent from municipal wastewater treatment plants (MWTPs), combined sewage overflows (CSOs), septic systems, and surface runoff. Emerging contaminants are defined here as pharmaceuticals and personal care products. Assessments of the presence, and ecological and human health risks associated with these contaminants in the Dnieper River’s surface waters have not been performed.

Data on the concentration and distribution of emerging contaminants in Ukraine is very limited. Vystavna et al. (2012) reported the presence of pharmaceuticals in riverine systems in eastern Ukraine (Kharkiv) and described the spatial, temporal and source characteristics of these compounds, however no information from the area around Kyiv or in the Dnieper exists. Pharmaceuticals have a broad distribution in aquatic systems (Kümmerer, 2009). Measured environmental concentrations of pharmaceutical compounds have been shown to cause adverse effects in aquatic organisms (Hoeger et al., 2005; Khan et al., 2018). In contrast to the emerging contaminants, there have been several studies addressing the distribution of legacy contaminants in selected Ukrainian riverine, estuarine and coastal regions. For example, metals (Burgess et al., 2011; Burgess et al., 2009; Linnik and Zubenko, 2000; Vystavna et al., 2012) and organic contaminants (Burgess et al., 2011; Burgess et al., 2009; Fillmann et al., 2002; Readman et al., 2002) have been reported over the last few years, however none performed in the Kyiv area.

The presence of contaminants does not necessarily indicate adverse ecological effects will occur. Toxicity testing is a tool to determine the bioavailability of contaminants in environmental samples and can integrate the complex chemistry of a milieu of contaminants. This allows environmental managers to prioritize areas with high toxicological signals for remediation or other regulatory actions. In addition to acute aquatic toxicity, specific endpoints such as mutagenicity allow us to assess chronic hazards of these undefined mixtures of chemicals. To assess adverse ecological effects, three assays were used to integrate the toxic signal from surface water samples. Test species were: 1) daphnids (Daphnia magna and Сeriodaphnia dubia), 2) the ciliate Colpoda steinnii, and 3) the bacterium Salmonella typhimurium (Ames assay). Each assay possesses unique strengths and sensitivities. The daphnids, D. magna and С. dubia toxicity tests are internationally well established assays for testing environmental water samples (Harmon et al., 2003; Knight and Waller, 1987; Versteeg et al., 1997). The protozoan C. steinii has been used in Ukraine and USSR as a toxicity testing organism (Pozdnyakova et al., 2006). C. steinii is an attractive test organism because it can be maintained in a cyst form and has a relatively short exposure time (10 min to 3 h). The Ames assay is a widely used and accepted test that assesses the mutagenic potential of chemical compounds (Ames and McCann, 1981; Ames et al., 1975; Zeiger, 1998). The S. typhimurium bacterial strains TA98 and TA100 are designed to detect frame-shift and base-pair mutations, respectively. The Ames assay is generally accepted to have a 65–90% concordance with mammalian carcinogenicity (Kirkland et al., 2005; Kitchin et al., 1992; McCann et al., 1988; Tennant et al., 1987). Because the bacteria are prokaryotes, S9 factor (an enzymatic mix derived from rat liver) is often added to mimic metabolism that may occur in a mammalian system. We used Strain TA98 and TA100, with and without S9 activation.

To better understand the presence, distribution and adverse biological effects of both legacy and emerging contaminants in the surface waters of Ukraine, we conducted a multi-year study focusing on seven surface water stations in the Dnieper River. Sampling was focused on stations located before and after the main Kyiv MWTP to determine the plant’s efficacy in removal of contaminants from surface waters. The goal of this investigation was to provide a novel survey of chemical concentrations of a set of known hazardous legacy chemicals as well as selected emerging contaminants and to examine the magnitude of mutagenicity and toxicity present in Kyiv’s section of the Dnieper River. The study provides initial, screening-level information on possible human health and ecological risks associated with Kyiv’s surface waters. Results of these chemical and toxicological analyses will inform scientists and environmental managers about the effectiveness of the MWTP and the condition of Kyiv’s surface waters.

2. Materials and Methods

2.1. Target Contaminants

To determine the concentrations and distribution of legacy and emerging contaminants, surface water samples were collected and analyzed for several contaminants including PCBs, PAHs, DDTs (including degradation products DDE and DDD), several other chlorinated pesticides, and six toxic metals. In addition, five classes of emerging contaminants (20 individual contaminants) were also monitored including stimulants, antibacterials/antifungals, hormones/estrogens, antibiotics, and pain relievers. For a full list of the 70 contaminants measured, see Supplementary Data Table S1.

2.2. Sampling Stations and Sample Collections

In the Fall of 2011 and Spring of 2012, 6.5 l of surface water samples were collected from up to seven stations (1 to 7) located near the Kyiv MWTP (Fig. 1, Supplementary Data Fig. S1). Station 1 is the influent into the MWTP and Station 2 is the effluent receiving ditch. Station 3 is the mouth of the effluent receiving ditch before it enters the Dnieper River. Station 4 is the Dnieper River at the mouth of the effluent receiving ditch. Station 5 is in the Dnieper River just downstream of the receiving ditch and Station 6 is in the Dnieper River upstream of the receiving ditch. Station 7 is the reference site in the Dnieper River upstream before the MWTP. Supplementary Data Table S2 provides further description of the sampling stations. In the Fall of 2013, another round of sampling was performed at a subset of stations (3, 4, 5, 7) to conduct a more intensive investigation of the bioavailable forms of contaminants by fractionating the water samples using a XAD column chromatography. Following the collections, samples were stored in coolers on ice until returned to the laboratory-generally within 24 h. Water samples were refrigerated at 4 °C in the dark until extracted and chemically analyzed. Water samples were split for various chemical analyses and data reported are splits from the same samples.

Figure 1.

Geographical location of sampling stations.

2.3. Sample Preparation and Chemical Analysis

2.3.1. Emerging Contaminants

Water samples (50 to 150 ml) for MWTP influent and effluent (station 1 and 2, respectively), and 400 to 1000 ml for other stations were filtered (0.45 μm cellulose filter), pH adjusted (~2.2 ± 0.05) and EDTA was added followed by solid phase extraction (SPE) with Supel Select 1000 mg HLB cartridges. Elutions were conducted using 6 ml 100% methanol and 3 ml methanol:acetone (1:1), the two fractions were combined, evaporated under gentle nitrogen stream and re-dissolved in 1.5 ml of methanol. Suspended matter samples were dried under a gentle nitrogen stream at ambient temperature and extracted by double ultrasonication using 100% methanol, and methanol:acetone (1:1) mixtures (4 m, 30 °C, 15 min). The extractions were then centrifuged (5 min, 3000 rpm), combined, evaporated and re-dissolved in 1.5 ml of methanol. Extracts were analyzed based upon EPA Method 1694 (U.S. Environmental Protection Agency, 2008) on an Agilent 1200 HPLC/MS Quadrupole 6130 in SIM mode with a Discovery HS C₁₈ 4.6 mm × 250 mm × 5 μm column at 30 °C with a gradient of water (phase A) and methanol:acetonitrile 50:50 (phase B) with 0.1% formic acid. Estriol, 17β-estradiol and estrone, were analyzed using precolumn derivatization with dansyl chloride in 0.1 M Na₂CO₃ buffer pH 9.65 (5 min, 60 °C), followed by 3 min centrifugation (3000 rpm) and immediate determination using a Zorbax Eclipse XDB-C₁₈ narrow bore (2.1 × 150 mm × 5 μm) column in the gradient water (phase A) and acetonitrile:water:methanol 85:10:5 (phase B) with 10 mM/l ammonium. Each compound was identified based on matching of retention times (± 0.01 min) and m/z values [M + H⁺] for all ECs except triclocarban and triclosan [M-H⁻] which were determined from analysis of standards (Pestanal®, Sigma-Aldrich®) and estrogen’s dansyl derivatizatives [M + H⁺]. Quantification was performed using five-point (5–500 ng/mL) calibration curves considering target substances recovery rates (85–112%). The limits of quantification (LOQs) were defined as the lowest concentration with a signal-to-noise ratio of 10 and varied from 0.01 to 0.07 μg/l for 1,7-dimethylxanthine, acetaminophen, caffeine, chlortetracycline, cortisol, cotinine, erythromycin, oxytetracycline, prednisolone, progesterone, sulfadimethoxine, sulfamethoxazole, testosterone, tetracycline, triclocarban, triclosan and trimethoprim. LOQs for estriol, 17β-estradiol and estrone were approximately 0.005 μg/l.

2.3.3. Legacy Metal Contaminants

Water samples were filtered (0.45 μm nylon filters) and acidified with 65% nitric acid (99.99% (American Chemical Society) 5 ml/l sample) before analysis. Divalent transition metals (Cd, Cu, Ni, Pb, Zn) were analyzed in duplicate using inductively coupled plasma mass spectrometry (ICP/MS), on an Agilent 7500 CE. Trace metal determination in water samples was performed according to the International Standards: ISO 17294–1:2004, and ISO 17294–2:2003 for 62 elements. In some cases, graphite furnace (Saturn Graphite-2, Chemavtomatika, Severodonetsk, Ukraine) or flame atomic absorption spectroscopy (AAS) (C-115-M1, Electron, Sumy, Ukraine) was used to confirm an analysis. Trace metal determination in water samples was performed according to International standards: ISO 15586:2003. For calibration, Interstate Standard solutions of ions of metals were used (Ukraine, Odessa, Physicochemical Institute). Pre-filtration resulted in two types of samples and concentration data is reported as total and <0.45 μm. For digestions, filter samples were dried (70 to 80 °C), and finely ground using an agate mortar. These samples were then processed using a microwave assisted digestion system (Speed Wave MWS-2, Berghof Products + Instruments GmbH) in closed digestion vessels containing HNO₃:HCl (3:1) for 30 min. Following extraction, digests were filtered through 0.45 μm polyvinylchloride membranes, diluted with deionized water to 10 ml, and stored at room temperature until analysis by ICP/MS. Method detection limits (MDLs) for the metals ranged from 0.01 to 0.1 μg/l.

2.3.3. Legacy Organic Contaminants

Following water sample filtration (1000 ml through a 0.45 μm High volume: Low Pressure (HVLP) Durapore membrane filter), samples were extracted using SPE polymeric (e.g., XAD-2) or C₁₈ sorbents with elution using nonionic organic solvents (acetone and methanol applied in sequence). Filter samples, following homogenization, were extracted by Soxlet extractor in 1:1 acetone:hexane for 16 h. All extracts were analyzed in duplicate. PCBs and chlorinated pesticides were analyzed by GC/MS on an Agilent GC 6890/MSD 5975I in SIM mode using an autosampler or Hewlett-Packard HP5890 series II with a single GC/ECD HP G1223 (only October 2013 samples). PAHs were analyzed by HPLC on a Waters Alliance with an E2695 separation module and 2998 photodiode Arrow 2475 multi wavelength fluorescent detector. The liquid chromatography was performed on a 250 × 4.6 mm column with Lichrosorb RP-18 sorbent (5 μm thick) using 50–100 μl volume injections with a combination of mobile phases: A (CH₃CN/H₂O, v/v 4:5) and B (CH₃CN) performed at 25 °C. MDLs for the PAHs were approximately 0.01 ng/l. Calibration of techniques for determination of OCPs, PCBs, PAHs was carried out on standard blends: Supelco EPA Pesticides Mix 48,858-U 10–60 μg/ml in MeOH:CH₂Cl₂ (98:2); Supelco EPA CEN PCBs Congener Mix 1 4–7927 10 μg/ml in heptane; Supelco EPA 610 PAHs Mix 4–8743 100–2000 μg/ml in MeOH:CH₂Cl₂ (1:1) respectively. To assess the precision and accuracy of the GC/ECD, GC/MS and HPLC analytical methods for OCPs, PCBs and PAHs, NIST standard reference materials (SRM) (1944 NIST) (New York/New Jersey Waterway Sediment) was analyzed. Precision and accuracy for this SRM ranged from 11.9–26.9% (mean = 17%) and 2.9–15.9% (mean 9.5%) respectively.

2.4. Toxicological Sample Preparation and Endpoints

2.4.1. Sample Preparation

Aquatic toxicity testing assays (e.g., Colpoda, Daphnia) were performed with whole water samples (i.e., no filtration, XAD or SPE). Mutagenicity assays were tested with whole extracts in Fall 2011 and Spring 2012. Preparation of the extracts for mutagenicity assays generally followed the procedures described above for the chemical analyses with the following changes. Whole extracts were prepared by passing water samples through a XAD-2 column with subsequent extraction by acetone and methanol. Mutagenic samples from Fall 2011 and Spring 2012 had concentration factors (reflecting the ratio of the original sample volume to 0.5 ml) ranging from 37 to 135×. In Fall 2013, in order to determine toxicity associated with particulate and dissolved phases, extracts were prepared using the following procedure: samples were filtered sequentially through a 24 μm filter (coarse), 0.45 μm filter (fine), and then a XAD-2 column to isolate dissolved phase contaminants. Particulates from both filters were combined, Soxhlet extracted with acetone:hexane (1:1), brought to dryness, and then re-dissolved in 0.5 ml of dimethyl sulfoxide (DMSO). The XAD-2 sorbent was extracted with acetone:hexane (1:1), brought to dryness, and then re-dissolved in 0.5 ml of DMSO. Sterile distilled water was added to 0.1 ml of the DMSO extract for a total volume of 20 ml. Samples from Fall 2013 had only a 20× concentration factor.

2.4.2. Ciliate Colpoda steinii Mobility Toxicity Test

This test was performed according to Pozdnyakova et al. (2006). Briefly, a dry culture of C. steinii was mixed with medium and incubated at 26 to 28 °С for 18 to 24 h. Immediately prior to the start of the study, the culture and the media were exposed to sunlight for 15 min to activate the culture and inspected under a light microscope to ensure that at least six motile cells of C. steinii were visible in 0.2 ml of culture. The 100% whole water sample and the culture were added in a 1:1 ratio (50% dilution factor) and incubated for up to 3 h. Samples were observed at 10 min and 3 h, and classified as extremely toxic, moderately toxic and non- toxic based on immobilization of C. steinii at each time point.

2.4.3. Daphnia magna and Сeriodaphnia dubia Acute Toxicity Tests

These assays were performed according to established methods (https://www.epa.gov/cwa-methods/whole-effluent-toxicity-methods). As noted, the organisms were exposed to 100% whole surface water. Briefly, organisms were cultured onsite at the Institute of Hydrobiology (Kyiv, Ukraine). For 96 h acute tests, 10 newly hatched (<24 h old) organisms were placed in each of three replicate exposure chambers. Animals were not fed and there was no water renewal during the acute toxicity tests. Test chambers were kept at 17 to 18 °C with ambient laboratory lighting (i.e., 16:8 light:dark). Observations of mortality were recorded at 24, 48, 72 and 96 h. The 96-h data are reported in this study.

2.4.4. C. dubia Chronic Generational Test

U.S. EPA protocols (U. S. Environmental Protection Agency, 2002) were followed to perform this method with the following modifications. Five replicate exposure chambers were used. Water was renewed three times during the 8 day exposure and organisms were fed the microalgae Clorella vulgaris. Test chambers were maintained at 18 ± 1 °C with ambient laboratory lighting (16:8 light:dark). Observations were carried out for three broods. The number of young/female was the endpoint used for comparison.

2.4.5. Ames Mutagenic Activity Assay

Mutagenicity was determined with a one-day old culture of the test strain of bacteria S typhimurium ТА98 or S typhimurium ТА100 with or without S9 using EPBI test kits (Ontario, Canada). Tests were performed according to EPBI instructions. Positive controls were 2-nitrofluorene (2-NF, 3.0 mg/l), a direct-acting mutagen, which does not require S-9 activation (for S. typhimurium ТА98), sodium azide (NaN3, 0.05 mg/l), a direct acting mutagen, which does not require S-9 activation (for S. typhimurium ТА100), and 2-amino-anthracene (2AA, 1.0 mg/l) in DMSO, an indirect-acting mutagen, which requires S-9 activation was used for both test strains. The negative control was “Morshinska”, a commercially-available Ukrainian bottled drinking water. As noted above, these assays were performed with three different fractions of the collected water samples: 1) coarse (24 μm), 2) fine (0.45 μm), and 3), dissolved phase (eluted from the XAD) as described above. Replication was not performed due to limited resources. The probability of detecting mutagenic and cytotoxic samples was determined using a probability rule at the 95, 99 and 99.9% confidence levels (Gilbert, 1980).

2.5. Statistical Analyses

Comparisons to detect significant differences among toxicological treatments were conducted using analysis of variance (ANOVA) followed by a Dunnett’s test to identify specific differences (SAS, Gary, NC, USA). Correlation analysis between water concentrations of contaminants and observed toxicity was performed using Excel (2016).

3. Results and Discussion

3.1. Concentrations of Emerging and Legacy Contaminants

3.1.1. Emerging and conventional contaminants in water samples

Fig. 2, Fig. 3, Fig. 4 report the concentrations of total dissolved emerging contaminants and metals, total PCBs and PAHs, and total DDTs and non-DDT pesticides, respectively, in water at stations 1 to 7. Emerging contaminants were dominated by stimulants, antibacterials/antifungals, antibiotics, and hormones/estrogens. Of the 20 individual emerging contaminants analyzed, the contaminants with the highest concentrations were triclosan (3 μg/l and 30 μg/g in the dissolved and particulate phases, respectively) and triclocarban (0.8 μg/l and 0.17 μg/g in the dissolved and particulate phases, respectively). Of the seven hormones analyzed cortisol (0.23 μg/l) and estriol (0.21 μg/l) had the highest concentrations and the rest were generally <0.1 μg/l. Caffeine (19.2 μg/l) had the highest concentration of the three stimulants measured. Acetaminophen was found at concentrations <0.7 μg/l, and the seven antibiotics measured generally had concentrations <0.4 μg/l. Supplementary Data Tables S3 to S11 provide more specific concentration information. Total dissolved influent concentrations of emerging contaminants (Station 1) ranged from 5.5 to 29 μg/L for Spring 2012 and Fall 2011 samples, respectively (Fig. 2a). In general, Fall 2011 emerging contaminant concentrations were approximately six times greater than the Spring concentrations. This may have been due to higher rainfall in Spring 2012 (215 mm) compared to Fall 2011 (100 mm) (rainfall data, Central Geophysical Observatory-Kyiv). The higher rainfall and runoff may have diluted the concentrations of emerging contaminants in surface waters.

Figure 2.

Dissolved concentrations (μg/L) of (a) total emerging contaminants (EC) and (b) total metals in surface waters collected in Fall 2011, Spring 2012 and Fall 2013 from stations around the Kyiv municipal wastewater treatment plant. Classes of ECs and individual metal contributions are also shown. Concentrations of emerging contaminants were below detection limits for Stations 5, 6 and 7. See supplemental information Tables S3 and S4 for specific EC and metal concentrations.

Figure 3.

Concentrations (ng/L) of (a) total PCBs and (b) total PAHs in surface waters collected in Fall 2011, Spring 2012 and Fall 2013 from stations around the Kyiv municipal wastewater treatment plant. See supplemental information Tables S5 and S6 for specific PCB and PAH concentrations.

Figure 4.

Concentrations (ng/L) of (a) total DDTs and (b) total non-DDT pesticides in surface waters collected in Fall 2011 and Spring 2012 from stations around the Kyiv municipal wastewater treatment plant. Individual pesticide contributions are also shown. See supplemental information Tables S7 and S8 for specific pesticide concentrations.

Effluent from the MWTP (Station 2) contained total concentrations of dissolved emerging contaminants of 0.12 and 1.3 μg/l for respective Spring and Fall collections. This represents a 96% reduction in the overall emerging contaminant load associated with the MWTP compared to the influent concentration at Station 1. The Kyiv MWTP has primary treatment which consists largely of mechanical removal through floatation and settling, and secondary biological treatment with activated sludge. No tertiary treatment (ultraviolet or chlorination) is performed. Removal of contaminants could occur through co-sedimentation or co-floatation with organic matter in the primary treatment, or degradation or co-sedimentation with surplus active sludge in the secondary treatment. Our results agreed with reported percent removal rates in a review of secondary treatment MWTPs in Spain (71–99%) for ECs that overlapped with this study (Gros et al., 2010). Concentrations of ECs at upstream and downstream riverine stations (Stations 5, 6, 7) (Fig. 2a) were below detection limits strongly indicating the primary source of this class of contaminants to the Dnieper River is the MWTP. Additional water samples collected at selected stations in Fall 2013 (Stations 3, 4, 5, 7) also had low concentrations of emerging contaminants at levels very close to the detection limits (Fig. 2, SD Table S9). Most ECs were found in the dissolved phase indicating little interaction with particles (SI Tables S3, S9). Exceptions to this observation were the antibacterial/antifungal triclosan and triclocarban which were both primarily associated with particulates (SI Tables S3, S9). Most of the ECs detected in this study had log K_OWs < 1.0 while triclosan and triclocarban have values of 4.76 and 4.90 (Perron et al., 2012; Trouts and Chin, 2015), respectively. As such, the expectation is that they will primarily partition to sediments where they may cause a higher exposure to sediment dwelling organisms and may biomagnify to water column organisms. This data indicates that wastewater is the primary source of ECs in the Dnieper River and also the important role of the MWTP in removal of ECs from wastewater before it enters the river. Peng et al. (2008) examined the concentration of four emerging contaminants in the Pearl River in southeastern China that overlapped with compounds investigated here. Concentrations of the four compounds ranged from 35 to 1023, non-detect (nd) to 50, nd to 2, and nd to 1 ng/l for triclosan, estrone, 17β-estradiol and estriol, respectively (Peng et al., 2008). In the Dnieper River, these compounds were found at much lower concentrations ranging from nd to 3.1, nd to 0.03, nd to 0.08, and nd to 0.21 ng/l for triclosan, estrone, 17β-estradiol and estriol (SD Table S3), respectively.

Dissolved total metals were found in the influent (Station 1) at concentrations of 150 and 39 μg/l for the Fall 2011 and Spring 2012 collections, respectively (Fig. 2b. Table S4). The Fall 2011 samples were about three times greater in metal concentrations with zinc, lead and copper dominating the distributions compared to Spring 2012. After passing through the MWTP, the concentrations of total metals were found to decrease by 53 to 64% at Station 2 indicating significant but not complete removal. Further, concentrations at the upstream and downstream stations showed dissolved metal concentrations of 22 to 26 μg/l for Fall 2011 samples collected from Stations 6 and 7 while Spring 2012 concentrations at those same Stations were 3.4 and 4.9 μg/L, respectively. Again, this may be related to higher rainfall during the Spring season diluting the metal concentrations. Unlike the emerging contaminants, these data suggest metals are being introduced into the river from sources in addition to the MWTP.

For this investigation, we focused primarily on total concentrations of most contaminants, however in Spring of 2012 and then 2013, we fractioned metal samples to gain information on metal partitioning between dissolved and particulate phases. In Spring 2012, water samples were filtered through a 0.45 μm nylon filter (SD Table S4). Both copper and lead were found to be associated with the particulate phase of samples while nickel, zinc and chromium were primarily (>75%) dissolved. These data suggest most copper and lead were not bioavailable (Di Toro et al., 1991). The Spring 2013 water samples were filtered even further into the 0.45 μm and 0.20 μm fractions (SD Table S10). In these samples, both nickel and cadmium were present primarily in the dissolved phases (i.e., <0.45 μm). Again, copper was found associated with the particulate phase. Zinc, lead and chromium demonstrated more variability in their distributions than observed in the Spring 2012 samples with inconsistent patterns between the dissolved phase (e.g., zinc at Station 3) and the particulate phase (e.g., zinc in Station 4). In an assessment of nine rivers feeding into Laizhou Bay in northern China, Xu et al. (2017) reported on the concentrations of the same metals investigated here. Concentrations ranged from 0.4 to 2800 μg/l, 0.4 to 2100 μg/l, 0.7 to 16 μg/l, 1.3 to 59 μg/l, 0.1 to 6.3 μg/l, and 0.9 to 3.3 μg/l for copper, zinc, chromium, nickel, cadmium and lead, respectively (Xu et al., 2017). In the Dnieper River, these metals demonstrated ranges of 0.3 to 14, not detected (nd) to 110, 0.4 to 5.8, 1.0 to 6.0, nd to 0.1, and nd to 24 μg/l for copper, zinc, chromium, nickel, cadmium and lead, respectively (pre-filtered values, SD Table S4). Except for copper and zinc the values are similar between the two river systems. In the Chinese rivers, both copper and zinc were more concentrated than in the Dnieper.

For organic contaminants, dissolved concentrations of total PCBs, total PAHs, total DDTs and total non-DDT pesticides ranged from 2.3 to 53, 0.51 to 306, 0.9 to 15.9, and non-detect to 22.5 ng/l, respectively (Fig. 3, Fig. 4). Unlike the ECs and metals, the total PCBs and total PAHs demonstrated higher concentrations in Spring 2012 compared to Fall 2011 ranging from four to 23 times greater and five to 41 times larger for the total PCBs and PAHs, respectively. Also, unlike the emerging contaminants, for total PCBs and PAHs, there’s limited evidence of any removal by the MWTP, as well as evidence for multiple sources of these compounds to surface waters. For example, for the total PCBs in Spring 2012, concentrations between Station 1 and 7 were observed to change relatively little (e.g., 53 to 27 ng/l) while Fall 2011 samples showed higher PCB concentrations in the Dnieper River than in the MWTP influent. For total PAHs, similar trends are observed with Fall 2011 water samples showing higher concentrations in the river than stations most associated with the MWTP (Stations 1 and 2, Fig. 3b). In addition, extra samples collected from selected stations in Fall 2013 showed total PAH concentrations of 0.5 ng/l at Station 3 climbing to 128 ng/l at Station 5 downstream of the MWTP.

Total PAHs measured in the Dnieper River water samples were dominated by phenanthrene, fluoranthene, and other molecules associated with petroleum sources (SD Table S6) (Burgess et al., 2003). This suggests petroleum products (e.g., oils) are leaking into the Dnieper River from multiple locations. Road runoff could be a possible major vector for these compounds to enter the river from spills on the terrestrial landscape. To make this conclusion more definitive water samples should be analyzed for alkyl PAHs (e.g., 1,3-dimethylnaphthalene, 2-methylanthracene, 3,6-dimethylphenanthrene) (U.S. Environmental Protection Agency, 2003). These alkylated molecules are diagnostic of PAHs originating from petrogenic sources as a result of the slow and incomplete degradation of ancient highly-alkylated plant material (Burgess et al., 2003).

After metals, PAHs were detected at the highest concentrations (up to hundreds of ng/l) in the Dnieper River surface waters (SD Tables S6 and S11). In general, hydrophobic organic contaminants, like PAHs, are considered most bioavailable in the dissolved form (Di Toro et al., 1991). In this investigation, PAHs associated with the coarse and fine fractions (SD Table S11) were operationally-defined as being associated with the particulate phase. For example, PAHs in Fall 2013 samples from Stations 3, 4 and 5 were associated principally with the particulate phase while only a small proportion of PAHs were associated with the XAD fraction (which captures dissolved chemical). Given the relatively high log K_OW of PAHs (e.g., 3.4 to 8.0) (U.S. Environmental Protection Agency, 2003) this distribution is expected. In contrast, at Station 7, most of the PAHs were associated with the dissolved phase. These findings suggest that while PAHs were often measured at elevated concentrations in the riverine surface waters (Fig. 3b), they may not have been bioavailable.

In summary, these data indicate that both PCBs and PAHs are present in the influent of the MWTP with little removal during the treatment process. In addition, the Dnieper River appears to contain elevated concentrations of these contaminants originating from other sources beyond the contribution from the MWTP. For comparison, Yu et al. (2018) surveyed the identical suite of PAHs measured in this study of the Dnieper River in surface and river waters near the Pearl River Delta. They found total PAHs ranging from 92.8 to 324 ng/l (Yu et al., 2018) which agree closely with measurements in the Dnieper River of 6.28 to 306 ng/L (SD Table S6).

Total DDTs distribution was similar to the PCBs and PAHs with higher concentrations in the Spring 2012 influent samples (i.e., approximately 18 times greater than Fall 2011 sample). Also, there was evidence the MWTP successfully removed 92% of the DDTs from the influent (Fig. 4a). However, water samples from the Dnieper River were often elevated compared to samples collected from near the MWTP (e.g., Station 4). In almost every case, the dominant DDT in the total DDTs mixture was 4,4′-DDT followed by 4,4′-DDD suggesting a relatively new source of parent DDT is present in the environment (Foght et al., 2001). According to Ukrainian environmental laws going back to the Soviet Union, DDT was banned for use in early 1970s as an agricultural pesticide (Melnikov and Shvindlerman, 1990), but practically eliminated by 1980. However, it is unclear how thoroughly this ban is enforced, and the presence of 4,4′-DDT suggests it continues to be used in the Kyiv area. In an earlier study, we found elevated levels of DDTs in the sediments of several Ukrainian estuaries (e.g., Sevastopol Bay, Dnieper and Boh estuaries, Danube delta) (Burgess et al., 2011). Historical use of DDTs would result in their presence buried deep in the sediments; however, they were found in the surface sediments suggesting recent usage. If DDTs have been used in the area around Kyiv until recently, the higher concentrations detected in the Spring 2012 water samples compared to the previous Fall may reflect seasonal application to agricultural lands. In contrast, at Station 7, Spring 2012 water sample is dominated by 4,4′-DDD possibly indicating a different source of pesticides to the northern section of the Dnieper River in Kyiv, a source containing a more thoroughly degraded DDT mixture (Foght et al., 2001). For comparison, a review of total DDTs in several environmental compartments in the Pearl River Delta including the Shenzhen, Dasha and Pearl Rivers reported a range of 0.03 to 190 ng/L (Guo et al., 2009) which was comparable, but greater than the range of total DDTs in the Dnieper River during our study (i.e., 0.90 to 15.8 ng/l (SD Table S7)).

None of the six non-DDT pesticides were detected in Fall 2011 water samples except at Station 7 (0.8 ng/l) (Fig. 4b). In Spring 2012 samples, the highest concentrations of the pesticide mixture were in Station 1 influent with 23 ng/L. The MWTP appeared to reduce the concentration of the mixture (~87% removal) but the concentrations of the pesticides in the water samples increased in the river with distance from the MWTP. Based on elevated concentrations of the pesticide mixture in the Dnieper River at Stations 5, 6 and 7, the presence of additional contamination sources from other locations than the MWTP is likely. None of the non-DDT pesticides was dominant although γ-hexachlorohexane (i.e., the pesticide Lindane) and hexachlorobenzene were commonly detected.

3.2. Toxicity and Mutagenicity

3.2.1. Ciliate Colpoda stenii Mobility Toxicity Test

The C. stenii control samples (bottled water) were non-toxic and test samples maintained pH values within acceptable parameters. Water samples were toxic to C. stenii from the MWTP influent (Station 1) in Spring 2012 samples, and where the drainage channel from the MWTP emptied into the mouth of the Dnieper River (Station 4) in Fall 2013 samples (Fig. 5). For this assay, all other stations at all other sampling periods were non-toxic. These non-toxic results may result from the insensitivity of the assay or may be due to the inadvertent 50% dilution of the sample by the culture water which occurs when testing protocols are followed. Of the two stations where the C. stennii assay responded with adverse effects, Station 1 (MWTP influent) is the station where other toxicity assays indicated positive results (see below).

Figure 5.

Acute toxicity measured by the D. magna, C dubia and C. stenni toxicity tests in water samples collected from the Dnieper River located in Kyiv, Ukraine. Results are (a) Fall 2011, (b) Spring 2012, and (c) Fall 2013. * indicates a significant difference from the control.

3.2.2. D. magna and C. dubia Acute Toxicity Tests

Control survival for all D. magna and C. dubia acute tests averaged 97 ± 5% and 96 ± 5%, respectively. Both D. magna and C. dubia indicated ≤10% toxicity in the upstream station (Station 7), and the station situated above the outfall of the receiving ditch from the MWTP (Station 6) for all three sampling events (e.g., Fall 2011, Spring 2012 and Fall 2013) (Fig. 5). The MWTP influent (Station 1) was the most toxic station, followed by the MWTP effluent (Station 2). Stations in the MWTP receiving ditch (Station 3) and just south of the receiving ditch in the Dnieper River (Station 4) had significant toxicity in the Fall 2013 samples. Other stations had moderate but not statistically significant toxicity ranging from 3 to 25% mortality. The acute toxicity appears to be associated with MWTP outfall as the riverine water was not toxic (Stations 5, 6, 7). Toxicants often associated with MWTPs include many emerging contaminants including antibacterial and antibiotic compounds (Perron et al., 2012). Because of their wide usage, triclosan and triclocarban are two of the more commonly monitored toxicants. Triclosan and trichlocarban are antibacterial and antifungal agents found in many soaps, cleaners, toothpaste and lotions (Perron et al., 2012). However, the concentrations of triclosan and triclocarban found in surface waters at Station 3 were <0.1 μg/l (Fig. 2a, SD Table S9), which is three orders of magnitude lower than literature values that researchers have reported for 48 h. C. dubia survival and reproduction (240 μg/l) (Tatarazako et al., 2004) and 48 h. D. magna EC50 (390 μg/l) (Orvos et al., 2002). In addition to triclosan and triclocarban, the other 18 emerging contaminants analyzed were generally not detectable (Fig. 2a, SD Tables S3, S9) at Station 3. The highest concentration of DDT and its metabolites was 15.9 ng/l reported at Station 1 in Spring 2012, which is two orders of magnitude lower than reported LC50s (1050 ng/l) in the literature (Mézin and Hale, 2004).

C. dubia Chronic Generational Test Chronic data from C. dubia exposed to whole water samples (Fig. 6) showed a significant decrease in brood size at Station 3 in all three sampling events (i.e., Fall 2011, Spring 2012 and Fall 2013). In Fall 2013, samples from Stations 3 and 4 showed significant acute toxicity as well as chronic toxicity to C. dubia. Station 3 was the location closest to the MWTP outfall for the chronic C. dubia assay. PAHs can affect Cereriodaphin sp. brood production (EC50s range from 11,000–271,000 ng/l); however, changes in offspring number generally occur at PAH concentrations higher than those observed in this study (i.e., ~300 ng/L) (Bragin et al., 2016). For example, the highest levels of PAHs in water samples where chronic toxicity sampling occurred (Fig. 3b, SD Tables S6, S11) showed no decreases in C. dubia brood production (e.g., Stations 4 (240 ng/l) and 5 (130 ng/l)). Total metal concentrations at Station 3 (Fig. 2b, Table SD Tables S4, S10) were also at least an order of magnitude lower compared to reported chronic EC50s (Cd 33.0, Cu 85.4, Zn 119 μg/l) (Naddy et al., 2015). Triclosan chronic tests performed on D. magna decreased the number of neonates/female at 64–128 μg/l (Peng et al., 2013), which is three orders of magnitude higher than the detected concentrations of 0.1 μg/l. As with all urban waters, particularly those associated with MWTP, many other toxic compounds exist in surface waters and are rarely analyzed and reported due to the complexity of the analysis and the overwhelming number of possible analytes. This analysis suggests that observed toxicity was the result of unmeasured contaminants or a mixture of contaminants.

Figure 6.

Chronic toxicity measured by the C. dubia three-brood assay in water samples collected from the Dnieper River located in Kyiv, Ukraine for Fall 2011, Spring 2012 and Fall 2013. * indicate a significant difference from the control.

No obvious patterns appeared in the toxicity data among years or seasons, but the toxicity is generally associated with the MWTP outfall. On the contrary, the chemistry results, in some cases, found higher concentrations of contaminants in the river water samples than in the MWTP effluent samples (e.g., PCBs, PAHs). But as noted regarding mixtures, there are many potential contaminants not monitored that may contribute to toxicity. In addition, these assays may not be particularly sensitive to measured compounds such as PCBs which are associated with long term, bioaccumulative effects. Finally, the D. magna and C. dubia assays responded similarly to the test samples from the different stations, while C. steinii appeared to be relatively insensitive overall.

3.2.3. Ames mutagenicity assay and data limitations

Mutagenicity data was reported from Fall 2011, Spring 2012 and Fall 2013 samples (Fig. 7a–j). However, only the data from the Fall 2013 had a consistent concentration factor (20×) and could be compared among stations, as well as among the coarse, fine and dissolved fractions, and across different bacterial strains and ± S9. The Fall 2011 and the Spring 2012 data (Fig. 6a–d) could only be compared between the +S9/−S9 of the same sample and between the TA100 and TA98 strains of the same sample as these were prepared with identical concentration factors. The Fall 2011 and Spring 2012 data could not be compared among the different stations or through time as they were prepared with different concentration factors (37–135×). Some samples displayed cytotoxicity (Fig. 6, yellow circles). This indicates that the exposure to the sample may have been too toxic for cell survival and therefore the cells did not have an opportunity to express any possible mutagenicity. Future research will need to investigate mutagenic effects in diluted samples that avoid cytotoxicity. Despite these limitations, these data are useful as a screen and give us a snapshot of mutagenicity in Kyiv’s riverine waters.

Figure 7.

Mutagenicity results measured by the Ames assay on water samples collected from the Dnieper River located in Kyiv, Ukraine. Results are (a) Fall 2011 whole concentrate TA98, (b) Fall 2011 whole concentrate TA100, (c) Spring 2012 whole concentrate TA98, (d) Spring 2012 whole concentrate TA100, (e) Fall 2013 coarse fraction TA98, (f) Fall 2013 coarse fraction TA100, (g) Fall 2013 fine fraction TA98, (h) Fall 2013 fine fraction TA100, (i) Fall 2013 dissolved TA98, and (j) Fall 2013 dissolved TA100.

3.3. General mutagenicity overall patterns and comparisons to PAH and metals

Mutagenicity or cytotoxicity in the Ames assay was observed at all stations in at least one sampling period (Fig. 7). This includes the stations located upstream of the MWTP. In addition, mutagenicity was not correlated with stations associated with the influent and effluent of the MWTP (Stations 1 or 2 and to a lesser degree station 3). This implies contaminants causing mutagenicity originated from non-point sources and may be more environmentally widespread relative to emerging contaminants like pharmaceuticals which are often found in higher concentrations associated with MWTPs. Concentrations of PAHs were also higher in stations away from the influence of the MWTP (Fig. 3b, SD Table S6, S11). This widespread contamination may include atmospheric deposition of PAHs from incomplete combustion of coal, oil and wood including vehicular exhaust, fossil fuel plant emissions, as well as creosotes, asphalt and petroleum spills (Abdel-Shafy and Mansour, 2016; Manzetti, 2013). The area surrounding the Dnieper River is largely urban and may have high levels of PAHs largely due to vehicular use (Van Metre et al., 2000). Again, PAH levels and Ames assay results appear to be relatively diffuse and not associated with MWTP outfall. Both the TA98 and TA100 strains are sensitive to PAHs. More specifically, PAHs listed in SD Table S1 and found in the Dnieper River could have been responsible for some of the mutagenicity observed. However, when mutagenicity results and PAH concentrations underwent regression analysis, the coefficients of determination (r²) with the highest values were generally relatively low <0.51. Some researchers have reported that Cd (13 μg/l) is the only commonly monitored metal (i.e., Cd, Cr, Cu, Hg, Ni and Zn) that causes a mutagenic response in the Ames assay (Codina et al., 1995); others have reported negative values for Cd but positive values for Co (120,000 μg/l) and Cr (250 μg/l) (Wong, 1988), or only Cr (Wung-Wai and Wai-Ping, 1981). While Cu and Zn are reported to enhance the mutagenicity of some contaminants (and when relatively abundant), they are not generally known as mutagenic by themselves (Fuji et al., 2016; Orešcanin et al., 2002; Sawai et al., 1998). Metal concentrations reported in the water column throughout all three sampling events ranged in the low μg/l (Fig. 2b, SD Tables S4, S10). While these levels may have contributed to the mutagenic pressure on the bacteria, they were at least an order of magnitude below the mutagenicity threshold reported in the literature.

3.3.1. Comparison amoung Fall 2013 samples: among all stations

In Fall 2013, four stations were sampled and tested with two bacterial strains (TA 98 and TA100) and two S9 variations (48 samples total). Mutagenicity and cytotoxicity incidence values were calculated by dividing the total number of positive mutagenic or cytotoxic samples by the total number of tested samples. Stations 4 and 5 had the highest incidence of mutagenicity (10%) and Station 7 had the next highest incidence of mutagenicity (6%). Station 3 displayed no mutagenicity however it did show cytotoxicity (12.5%) which may have masked mutagenicity. This general pattern of widespread mutagenicity that is not closely associated with the MWTP is consistent with the observed mutagenicity and PAH concentration patterns in other years.

3.3.2. Comparison among Fall 2013 Samples: TA98 vs TA100

In the Fall 2013 samples, the TA98 strain (frame-shift mutation) has an overall mutation rate of 20% and appears to be more sensitive to mutagenicity than the TA100 (base-pair substitution) which did not detect any mutagenicity. The TA100 strain appears to be more sensitive to cytotoxicity (33% toxicity) relative to TA98 (8% cytotoxicity). This has also been observed in another study (Jurado et al., 1993).

3.3.3. Comparison among Fall 2013 samples: -S9 vs +S9

Samples without S9 (direct acting) had a higher incidence of mutagenicity in both the TA98 (42%) and the TA100 (42%) samples relative to the +S9 samples for both bacterial strains (0%). This would imply the mutagenicity or cytotoxicity is direct-acting and does not have to be induced by enzymatic systems. Direct acting mutagens include nitroarenes, nitroaromatics and oxy-substituted PAHs (Ball et al., 1990; Lemieux et al., 2009; Pederson and Siak, 1981; Salmeen et al., 1984). Many nitro-aromatic compounds are anthropogenically produced as precursors in dyes, pesticides, pharmaceuticals and munitions (Ju and Parales, 2010). Also, nitro-pyrenes are known mutagens and constituents of incomplete diesel combustion (Purohit and Basu, 2000). These contaminants are common environmental pollutants that could, in part, be responsible for some of the observed mutagenicity.

3.3.4. Comparison among Fall 2013 samples: size fractions

As discussed, to better understand contaminants bioavailability, the 2013 samples were fractionated into coarse, fine and dissolved aliquots. There appeared to be no differences or patterns for mutagenicity or cytotoxicity rates among these different aliquots. This implies that the cytotoxic or mutagenic compounds were not associated with any size fraction, and there was not a single active agent. Instead, several contaminants that are biologically active and associated with different particle size classes or dissolved contaminants were likely responsible for the observed mutagenic toxicity. While examining the mutagenicity of the divided sample gives information on the toxins associated with that size class, it may also have diluted the overall toxic/mutagenic potential of the sample, as well as removed any synergistic interactions that may have occurred.

3.3.5. Comparison between + and −S9 and strains at the same Stations 2011 and 2012

For the Fall 2011 and Spring 2012 results, comparisons can only be performed between the −S9 and the +S9, and between the TA98 and TA100 samples at the same stations due to differing concentration factors at different stations. For these two sampling periods, the TA98 strain detected more mutagenicity than the TA100 strain, and the TA100 strain appeared to be more sensitive to cytotoxicity, as noted previously in the 2013 samples and Jurado et al. (1993).

The assay also responded differently to −S9 and the +S9 samples for both the TA98 and the TA100 strains over 60% of the time, but no pattern emerged with respect to bacterial strain or the presence/absence of S9. In 50% of the samples, the bacteria responded more sensitively to either cyto-toxicity or mutagenicity with +S9, but in 13% of the test samples, the −S9 samples were more sensitive to mutagenicity with no differences in toxicity.

3.3.6. Comparison to other areas

Of the 96 samples evaluated for mutagenicity in our study (i.e., differing sites, seasonal sampling and +/− S9 combinations), 18 samples (19% of all samples tested) were mutagenic and 24 samples (25% of all samples tested) were cytotoxic. This percentage of mutagenicity in environmental samples is within the range of many other studies; however, comparisons among mutagenicity studies is often difficult to interpret as researchers often prepare their samples using different methods of collection and concentration, and sampling sites are often targeted rather than selected randomly. At river sites influenced by petrochemical industries in Brazil, Vargas et al. (1993) found 34% of all samples tested had positive results with strains TA98 and TA100. Of the sites that were positive, 6% were upstream of petrochemical influence, 82% were in the zone of influence of the petrochemical plants, and 12% were downstream of the petrochemical site. The TA98 bacterial strain without S9 was the most sensitive. Their results show a low to moderate level of mutagenicity even outside the zone of petrochemical influence. In the Rhine River (Germany), mutagenicity was found at all seven sites sampled primarily with TA98 with S9 activation (Hendriks et al., 1994). However, no mutagenicity was detected in their one reference site in the Markermeer River. Other researchers confirmed the mutagenicity of all tested Rhine River samples using TA98 activated with S9 (Slooff and Van Kreijl, 1982). In the River Seine, in France, of the 36 samples tested 28% were positive with the majority of the positive samples resulting from S9 activated TA98 samples (Vincent-Hubert et al., 2017). We also found the TA98 strain to be more sensitive to environmental river mutagenicity than the TA100 strain.

Conclusions

Measurements of emerging and legacy contaminants in water samples from seven stations near the MWTP and receiving waters of the Dnieper River demonstrate exposures to aquatic life were present. Some pollutants, like the emerging contaminants were present in the MWTP influent but were not detected in the post-treatment plant effluent or in the Dnieper River. Other contaminants were removed to varying extents from the influent by the treatment plant including some of the toxic metals and pesticides. In contrast, PCBs and PAHs were not effectively removed by the MWTP. In addition, the Dnieper River is contaminated by pesticides, PCBs and PAHs from other sources in addition to the MWTP. These data suggest that organisms affected by the Kyiv MWTP and living in the nearby Dnieper River are exposed to a variety of toxic legacy chemicals.

Levels of toxicity and mutagenicity measured with a variety of assays indicated that contaminants in the surface waters of the Dnieper River were bioavailable. While acute and chronic toxicity was associated with the influent and effluent of the MWTP, mutagenicity was not. Mutagenicity was noted in surface waters at least once at every station sampled including those upstream of the MWTP. As noted, relatively high levels of PAHs found at stations upstream of the MWTP outfall do not appear to be associated with the MWTP. This implies that mutagens may be dispersed throughout the Dnieper River via non-point sources including environmental deposition from fossil fuel combustion and other processes. This investigation is the first survey to examine the magnitude of mutagenicity and toxicity present in Kyiv’s Dnieper River along with extensive monitoring of chemical concentrations. This novel survey provides initial screening level information on possible human health and ecological risks associated with surface waters. More research on contaminant, toxicity and mutagenicity seasonality, sources and differences among stations would assist in completing a comprehensive risk assessment of the Dnieper River surface waters.