Article pubs.acs.org/est
Eurasian Dipper Eggs Indicate Elevated Organohalogenated Contaminants in Urban Rivers Christy A. Morrissey,*,†,‡ David W. G. Stanton,‡ M. Glória Pereira,§ Jason Newton,⊥ Isabelle Durance,‡ Charles R. Tyler,¶ and Steve J. Ormerod‡ †
Department of Biology and School of Environment and Sustainability, University of Saskatchewan, 112 Science Place, Saskatoon, Saskatchewan S7N 5E2, Canada ‡ Catchment Research Group, School of Biosciences, Cardiff University, Cardiff CF10 3AX, United Kingdom § NERC Centre for Ecology and Hydrology, Lancaster Environment Centre, Library Avenue, Bailrigg, Lancaster LA1 4AP, United Kingdom ⊥ NERC Life Science Mass Spectrometry Facility, Scottish Universities Environmental Research Centre, Rankine Avenue, East Kilbride G75 0QF, United Kingdom ¶ Biosciences, Geoffrey Pope Building Laboratories, University of Exeter, Exeter, Devon EX4 4PS, United Kingdom S Supporting Information *
ABSTRACT: Many urban European streams are recovering from industrial, mining, and sewage pollution during the 20th century. However, associated recolonization by clean water organisms can potentially result in exposure to legacy or novel toxic pollutants that persist in the environment. Between 2008 and 2010, we sampled eggs of a river passerine, the Eurasian dipper (Cinclus cinclus), from 33 rivers in South Wales and the English borders (UK) which varied in catchment land use from rural to highly urbanized. Dipper egg δ15N and δ13C stable isotopes were enriched from urban rivers while δ34S was strongly depleted, effectively discriminating their urban or rural origins at thresholds of 10% urban land cover or 1000 people/km2. Concentrations of total polychlorinated biphenyls (PCBs) and polybrominated biphenyl ethers (PBDEs) were positively related to urban land cover and human population density while legacy organochlorine pesticides such as p,p′-DDE, lindane, and hexachlorobenzene were found in higher concentrations at rural sites. Levels of PBDEs in urban dipper eggs (range of 136−9299 ng/g lw) were among the highest ever reported in passerines, and some egg contaminants were at or approaching levels sufficient for adverse effects on avian development. With the exception of dieldrin, our data shows PCBs and other organochlorine pesticides have remained stable or increased in the past 20 years in dipper eggs, despite discontinued use.
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INTRODUCTION Two major issues impacting freshwater biodiversity loss are global climate change1 and land use change.2 Urban development has become a particularly significant issue given that more than half of the world’s human population now lives in urban areas with a projected 68% (6.3 billion people) expected to live in cities by 2050.3 While the total area converted to urban land use is relatively small, biological impacts are disproportionately large.4 Transforming land cover toward urbanization is increasingly pervasive, for example, through physical, chemical, and biological changes in receiving rivers. Numerous studies have confirmed biodiversity impairment at virtually every trophic level in rivers draining urban areas.4−6 Urban effects on downstream water quality are often large and reflect pollution from diffuse and point-sources such as mining, industry, road-runoff, treated wastewater, stormwater drainage, and combined sewage overflows that discharge during heavy rainfall.7−9 Urban streams and rivers often contain complex mixtures of nutrients, pathogens, and organic carbon along with toxic substances such as dioxins, polychlorinated © 2013 American Chemical Society
biphenyls (PCBs), polyaromatic hydrocarbons (PAHs), polybrominated diphenyl ethers (PBDEs) and organophosphate flame retardants (OPFRs), phthalates, alkylphenols, pharmaceuticals and personal care products, natural and synthetic hormones, heavy metals and numerous pesticides.10−14 Most standard treatment practices in Europe and North America are not effective at removing many of these substances which sometimes results in adverse health impacts on fish and other aquatic organisms.15,16 In contrast to the rich literature on aquatic fauna, there is only limited understanding of the risks to riparian endotherms such as aquatic birds. This is despite indications that birds can be affected negatively by urban land use,17−19 and birds in industrial or urban environments often contain greater burdens of PCBs, dioxins, and/or PBDEs than elsewhere.20−22 Received: Revised: Accepted: Published: 8931
May 13, 2013 June 28, 2013 July 2, 2013 July 2, 2013 dx.doi.org/10.1021/es402124z | Environ. Sci. Technol. 2013, 47, 8931−8939
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In the United Kingdom, as the first industrialized nation, there has been a long history of river pollution23 as well as legislation intended to improve water quality. Severe pollution problems from coal mining, associated industries, and underperforming wastewater treatment have been progressively controlled over the past 20 years allowing clean water species to recolonize rivers.23,24 Paradoxically, however, such recolonization has created new ecotoxicological challenges where organisms are exposed to legacy pollutants or toxic substances that persist either in sediments or through conventional treatment processes.10,14,25 Riparian organisms include mammals such as otters (Lutra lutra) and river birds such as kingfishers (Alcedo atthis) or Eurasian dippers (Cinclus cinclus) that are not considered under the current legislation (e.g., Water Framework Directive 2000/60/EC) intended to limit priority toxic substances in surface waters.26 Key questions arise, therefore, about (i) how well the risks to such organisms are now being controlled and (ii) whether such organisms might indicate wider environmental risks arising from urban pollution. The need to understand pollutant effects in urban areas imply also the need for tools that help to predict where risks are greatest. For example, there is interest in land-use map-based thresholds of urbanization at which impaired stream conditions occur.27,28 Definitions used to classify “urban” areas are broad but typically describe areas developed for residential, commercial, industrial, and transportation purposes. Urban land is defined often using human population density or using the proportion of land covered by constructed (impervious) materials.4 Negative effects on stream indicator organisms often increase with urban land use descriptors such as percentage of urban land in the catchment, total impervious surface area (TIA: proportion covered by surfaces such as roofs and roads), effective impervious areas (EIA: surface area directly connected to streams), or road density.27 However, debate exists over whether there is a consistent threshold value for detecting urban impacts. For example, King et al.29 determined a threshold for urban effects on invertebrate assemblages between 21 and 32% urban development. Others have detected similar effects at lower urban cover between 8 and 12% TIA30 or 2.5 and 15% TIA.31 Effects on fish species presence and abundance can occur between 2 and 4%32 or 6−11% EIA.33 Regardless, there is little information on whether such thresholds of urban land cover can be applied to pollutant exposure of biota and, in particular, riparian vertebrates. Coupled with map-based predictors, management tools that detect urban inputs to river food-webs would also help to predict pollutant exposure risk. Specifically, previous work has shown how stable isotope tracers can indicate urban discharge and help to diagnose their effects on individual organisms or communities.34,35 Isotope ratios of carbon, nitrogen, and sulfur are frequently dominated by the heavier isotope in wastewater and industrial discharges and are conserved through freshwater food webs, thereby acting as tracers of organic pollution.36−39 Predictable changes in stable isotope ratios have been observed in aquatic organisms in catchments with increased human population densities and wastewater seepage from septic systems.38,40−42 However, relating these changes in stable isotopes to toxic urban pollutants in biota is rare. This study investigates stable isotopes and contaminant residues in eggs of a recognized bioindicator species, the Eurasian dipper (Cinclus cinclus), to assess the magnitude of urban stream pollution and determine threshold percentage
land use or population densities that could indicate adverse pollutant exposure. Working in some of the most historically polluted river systems in the whole of Europe, the South Wales valleys of the United Kingdom,43 our specific objectives were to (1) determine the utility of stable isotope profiles in dipper eggs (δ13C, δ15N, and δ34S) as a means of classifying streams and verifying thresholds of urban land use at which contamination in eggs occurs; (2) evaluate the relationships between persistent organohalogenated contaminants in eggs and urban stream indicators (% total impervious area and human population density and stable isotope tracers), and (3) assess the degree of current persistent organic pollution in urban rivers of south Wales relative to historic data.
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EXPERIMENTAL SECTION Background on the Study Area. We specifically chose South Wales and the mid Wales/English borders for this project because of its pollution history, the availability of background data, and a long-standing research interest in dippers and other river birds exists.44,45 During the 1970s, over 60% of South Wales’ rivers in the Taff, Rhondda, Cynon, Rhymney, Ely, Ebbw, Ogmore, and Clydach systems were grossly polluted by discharge from poorly performing wastewater treatment works and leaking trunk sewers.43 Biochemical oxygen demands were high (>10−15 mg/L), concentrations of ammonia exceeded 2−3 mg/L, and dissolved oxygen concentrations were depressed often below 40−60% saturation. Effluents from coal and coking plants added high concentrations of inert solids along with toxic pollutants including phenols and cyanide.43 Since then, extensive recovery has followed a combination of industrial decline and improved regulation, allowing recolonization of typical clean water organisms.24 Nevertheless, many combined sewage overflows still discharge untreated sewage at high river flow, while the presence of pharmaceuticals, personal-care products, and other complex organic compounds in surface waters indicate these substances can persist through conventional treatment.9 Further north in Wales and its eastern border with England, the Usk and Wye river systems drain largely from farmland and seminatural habitats and provide an important non-urban contrast with South Wales.46 Field Methods. From 2008 to 2010, we located Eurasian dipper nests and sampled in total 74 eggs (n = 69 eggs for contaminant analyses), from 33 different rivers in South Wales and the Welsh-England borders. Nests were followed through nest building to laying, and a single random egg from either the first or second clutch (avoiding duplicate samples from the same nest) was taken under authority during the first 7−10 days of incubation. All eggs were candled to determine fertility at the time of egg collection. While the majority (90%) of eggs were fresh and viable, we also opportunistically included abandoned or addled eggs from nests not previously sampled for stable isotope analysis only. All eggs were transported on ice and frozen in the shell at −20 °C. Egg contents were transferred into chemically clean jars within 6 weeks and stored frozen at −80 °C until ready for analysis. Human Population Density and Land Use Spatial Analyses. Human population densities for England and Wales were based on the 2001 and 2011 UK Census statistics obtained from downloadable spatial files (Office for National Statistics http://www.ons.gov.uk/ons/guide-method/census/ 2011/index.html) containing digital geographical boundaries and population densities at the lowest available resolution 8932
dx.doi.org/10.1021/es402124z | Environ. Sci. Technol. 2013, 47, 8931−8939
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Data Analysis and Mapping. We produced maps highlighting major zones of human population densities in Wales and the English borders based on census data mapped by district authority.49 Major rivers in Wales were overlaid,50 and dipper egg sampling locations were sized by relative contaminant concentrations (ng/g lipid weight) for sum PCBs, sum PBDEs, and p,p′-DDE. We verified assumptions of normality, homogeneity of variances, and noncolinearity between the variables.51 Data exploration revealed the variable urban land use (%) violated assumptions of normality and heterogeneity of variance which was improved after square root transformation (x + 1). Similarly, contaminant data were log-transformed to approximate normality and to ensure bivariate normal distributions for multivariate analyses.52 Concentrations which fell below detection were given a value of half the detection limit for summary statistics, multivariate analyses, and modeling purposes. Contaminant data were statistically analyzed using lipid normalized concentrations but further summarized as wet weights with % lipid for comparison with other published studies (Table S2, Supporting Information). An initial assessment included correlations between individual stable isotope ratios (δ15N, δ13C, and δ34S) and map-based measures of urban influence (% urban land use, population density). We evaluated whether derived axes from principle components analysis (PCA) of combined egg stable isotope measurements δ15N, δ13C, and δ34S could be used as an indicator of urban influence. A PCA of the 3 isotope ratios (centered and standardized) reduced the dimensions and variation into 2 axes. PCA factors 1 and 2 were each regressed against map-based measures of urban influence. Hierarchical clustering techniques were further applied to evaluate the structure of the samples such that sample eggs with similar isotope signatures clustered more closely together and the final clustering sequence was visualized using a dendrogram. We applied linear discriminant analysis to test formally whether eggs were correctly clustered using predefined categories for “urban” and “rural” locations. Eggs were defined as “urban” at human population density ≥1000 people or ≥10% urban land use within a 1 km2 buffer area around the dipper nest site. These values approximated the median thresholds frequently cited in other studies of urban impacts.28,35,53,54 We assessed each potential indicator of urban influence (% urban, population density, δ13C, δ15N, and δ34S, 3 isotope PCA factor scores) to determine their correlations with contaminant concentrations in matching eggs (sum PCBs, sum PBDEs). Multiple regression models and model selection methods assessed which factors explained the greatest variation in egg contaminant concentrations, and the most parsimonious models based on Akaike’s Information Criterion adjusted for small sample sizes (AICc) are reported.55 Finally, a PCA of the contaminant congeners (PCBs and PBDEs) was used to fingerprint patterns that may be associated with urban sources. Contaminant congeners (log10 transformed) were included in the PCA if they were detected in more than 65% of samples. All statistical analyses were conducted using JMP v.10 software (SAS Institute).
(Lower Super Output Areas, areas based on an average of 1100 inhabitants). Land use was identified using the Countryside Council for Wales Habitat phase 1 survey data, which provides broad characterization of vegetation communities and land-use pattern.47 The percentage of urban land use is equivalent to % total impervious area (TIA) including all paved and built up areas around each nest site. A 1 and 3 km2 buffer around the GPS location of each dipper nest site was delimited using ArcGIS v.10.1. Human population density and land use within the buffers was calculated using Geospatial Modeling Environment software (http://www.spatialecology.com/gme/). No significant differences were apparent in the estimated population density and % urban land cover using the two buffer sizes. Therefore, we selected the 1 km2 buffer as an index of the relative land use impacts around the nest site which typically captures the foraging area within a dipper territory.48 Chemical Analysis. Egg samples were analyzed at the Centre for Ecology and Hydrology in Lancaster, UK for organic chemical contaminants (organochlorine pesticides: p,p′-DDT, DDE, and DDD, dieldrin, α and γ- hexachlorocyclohexane (HCH), hexachlorobenzene (HCB); 36 PCB congeners and 23 PBDE congeners) (Table S1, Supporting Information). Once thawed, subsamples of each homogenized egg (∼1 g) were weighed accurately, ground with sand, and dried with anhydrous sodium sulfate. Each sample was Soxhlet-extracted in DCM for 16 h. A small portion of the extract was evaporated to zero volume, and the lipid content was determined gravimetrically. The remaining extract was cleaned using automated size exclusion chromatography followed by 5% deactivated alumina. The extract was finally divided into two fractions: one was spiked with labeled internal recovery standards 13C organochlorines (OCs) or 13C PCBs, and the other fraction was spiked with 13C PBDEs. A 20 μL subsample of each extract was injected into a GC-MS with the programmable temperature vaporization (PTV) inlet using two different methods for OC/PCBs and PBDEs. The entire GC-MS process used a 50 m (OCs and PCBs) or 25 m (PBDEs) HT8 column (0.22 mm internal diameter and 0.25 μm film thickness, SGE Milton Keynes, UK) with helium as the carrier gas (2.0 mL min−1). Residues were quantified using an internal standard method and calibration curves of commercially available standards for OCs and PCB and PBDE congeners (Greyhound Ltd., Birkenhead, UK and LGC Ltd., Teddington, UK). Procedural blanks were run concurrently, and all samples were recovery-corrected. Detection limits averaged 0.1−3.1 ng/g ww for most congeners and compounds measured (Table S1, Supporting Information). Stable Isotope Analysis. A 1 mL subsample from each egg homogenate (n = 69 from the same eggs used for contaminant analysis) was freeze-dried for stable isotope analysis. Approximately 1 mg of each dry sample was weighed into tin capsules and analyzed for δ13C and δ15N at the NERC Life Science Mass Spectrometry Facility, East Kilbride, Scotland. Approximately 2 mg of the same samples were packed in separate tin capsules and analyzed for δ34S analysis at IsoAnalytical, Crewe, UK. Carbon and nitrogen isotopes were analyzed in a Europa Hydra 20/20 continuous flow isotope mass spectrometer (CF-IRMS), and sulfur isotope analysis used elemental analysis-isotope ratio mass spectrometry (EA-IRMS). Three in-house standards were run every 10−12 samples for quality assurance. Results were reported in δ notation as the deviation from standards in parts per thousand (‰). All standard measurement errors were within 0.2‰.
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RESULTS Urban Indicators. All 3 stable isotopes in eggs were correlated with catchment land use. Both the δ13C (r = 0.37, p = 0.001) and δ15N (r = 0.26, p = 0.027) signatures in eggs increased with urbanization while δ34S (r = −0.63, p < 0.0001) 8933
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Eighty-four percent (54/64) of egg samples were classified correctly using a predetermined cutoff rule for urban sites as ≥10% impervious area or ≥1000 people/km2. Thus, this threshold could be generally applied, and the combined isotopic measures were used to infer dipper exposure to urban pollutants. Relationship between Contaminants and Urban Indicators. Contaminant concentrations varied widely among streams (Appendix 1), but large scale spatial patterns were clearly linked to varying human population density (Figure 2a−c). In particular, the industrial pollutants PCBs and PBDEs were more uniformly elevated in urbanized regions with higher human densities (Figure 2a,b). They were also strongly intercorrelated (r = 0.62, p = 0.0001) suggesting similar sources. Both PBDE (t67 = 5.08, p < 0.0001) and PCB (t67 = 3.66, p = 0.0013) concentrations were higher on urban streams than rural streams (Table 1). Model selection incorporating all potential indicators of urban influence revealed egg PCB concentrations were best predicted by % urban land use (p < 0.0001), δ15N (p = 0.004), and δ34S (p = 0.048) (reduced model: F3,55 = 13.8, r2 = 0.43, p < 0.0001, AICc = 62.2). Similarly, the most parsimonious PBDE model included % urban land use (p = 0.0002) and δ 15N (p = 0.0003) and δ34S (p = 0.21) (reduced model: F3,55 = 24.1, r2 = 0.57, p < 0.0001, AICc = 64.5) (Figure 3). PCA of log10 PBDE congener patterns and log10PCBs revealed clear shifts with increasing urbanization (Figure S1a,b, Supporting Information). The first 2 axes explained 83% and 78% of the variation in PBDE and PCB congeners, respectively. BDE congeners 66, 85, 99, 118, 153, and 126 and PCB congeners 153, 138, 180, 170, 163, 187, and 128 were most prevalent in urban dipper eggs which all strongly influenced the loadings. PCAPBDE sample scores on axis 1 (r2 = 0.46, p < 0.0001) were positively related to % urban land cover. Similarly, PCAPCB sample scores on axis 1 (r2 = 0.53, p < 0.0001) and axis 2 (r2 = −0.14, p = 0.005) were each related to % urban land cover. Consistent with historic agricultural use, the legacy organochlorine pesticides such as p,p′-DDE and dieldrin were more
Figure 1. Principle component analysis biplot showing sample scores and relative weightings of 3 stable isotope ratios (δ13C, δ15N, and δ34S) measured in dipper eggs from predefined urban sites (filled symbols: ≥10% urban or ≥1000 people/km2) or rural sites (open symbols:
