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Arsenic in stream waters is bioaccumulated but neither biomagnified through food webs nor biodispersed to land
Ecotoxicology and Environmental Safety 139 (2017) 132–138
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Arsenic in stream waters is bioaccumulated but neither biomagnified through food webs nor biodispersed to land
MARK
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Luiz U. Heppa, , João A.M.S. Pratasb, Manuel A.S. Graçac a b c
Departament of Biological Sciences, Universidade Regional Integrada do Alto Uruguai e das Missões, 99709-910 Rio Grande do Sul, Brazil MARE – Marine and Environmental Sciences Centre, Department of Earth Sciences, University of Coimbra, 3001-517 Coimbra, Portugal MARE – Marine and Environmental Sciences Centre, Department of Life Sciences, University of Coimbra, 3001-401 Coimbra, Portugal
A R T I C L E I N F O
A B S T R A C T
Keywords: Biomagnification Isotopic approach Arsenic contamination Macroinvertebrates Water quality
Human activities such as mining have contributed substantially to the increase of metals in aquatic environments worldwide. These metals are bioaccumulated by aquatic organisms and can be biomagnified along trophic webs. The dispersal of contaminants from water to land has been little investigated, even though most aquatic invertebrates in streams have aerial stages. We used field and laboratory approaches to investigate the effects of arsenic pollution on stream invertebrate assemblages, and its bioaccumulation, biomagnification and trophic transfer from aquatic to terrestrial environments by emergent insects. We conducted the study in an arsenicimpacted stream (40 μg L–1 As at the most polluted site) and a reference stream (0.3 μg L–1 As). Invertebrate abundance and richness were lowest at the most impacted site. Arsenic in biofilm and in invertebrates increased with the arsenic content in the water. The highest arsenic accumulators were bryophytes (1760 μg g–1), followed by the biofilm (449 μg g–1) and shredder invertebrates (313 μg g–1); predators had the lowest arsenic concentration. Insects emerging from water and spiders along streambanks sampled from the reference and the impacted stream did not differ in their body arsenic concentrations. In the laboratory, the shredder Sericostoma vittatum had reduced feeding rates when exposed to water from the impacted stream in comparison with the reference stream (15.6 vs. 19.0 mg leaves mg body mass–1 day–1; p < 0.05), but they grew faster in the polluted water (0.16 ± 0.04% day–1 vs. 0.05 ± 0.01% day–1, p < 0.05). S. vittatum exposed to contaminated stream water accumulated arsenic from food, not through contact with water. We concluded that although arsenic is bioaccumulated, mainly by food ingestion, it is not biomagnified through food webs and is not transported from the aquatic to terrestrial environment when insects leave the stream water.
1. Introduction As a consequence of human activities, many natural and artificial compounds are being released into the environment. Many such compounds enter the food webs and profoundly harm the environment and humans. Some pollutants are accumulated by living organisms, particularly by primary producers (Ancion et al., 2010; Varun et al., 2012; Vithanage et al., 2011). When the concentration of pollutants is higher in the organisms than in their environment, there is a possibility of biomagnification through food webs, particularly affecting higher trophic levels. If highly mobile organisms bioaccumulate pollutants, they can disperse these pollutants away from the source of contamination. Most invertebrates in low-order streams are insects with an aerial stage, and therefore capable of transporting pollutants to land (Mogren et al., 2013). We investigated the potential movement of a pollutant
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from water to land by insects, using arsenic as the model. Metals and metalloids are common in nature, but can be toxic even at low concentrations (Gall et al., 2015). Biomagnification of Cd, Zn, Hg and Se was observed in freshwaters, but the mobility of other elements through food webs remains unclear (Cui et al., 2011; Ikemoto et al., 2008; Juncos et al., 2016; Kehrig et al., 2013; Painter et al., 2016). Arsenic is a toxic trace element that occurs naturally in the environment. Arsenic in water may originate from rocks by weathering (Alonso et al., 2014; Wang and Mulligan, 2006) and volcanic activity (Juncos et al., 2016). Human activities, including industry, municipal wastes and particularly mining, are important sources of arsenic to the environment (Liu et al., 2010; Subhani et al., 2015). Fungicides, herbicides, insecticides and phosphate fertilizers are also potential sources of arsenic contamination (Ghaeni et al., 2015; Wang and Mulligan, 2006; Wang et al., 2015). All organisms can take up arsenic (as arsenite) because it is chemically similar to phosphate (Wang and
Correspondence author. E-mail address: [email protected] (L.U. Hepp).
http://dx.doi.org/10.1016/j.ecoenv.2017.01.035 Received 28 September 2016; Received in revised form 16 January 2017; Accepted 19 January 2017 0147-6513/ © 2017 Elsevier Inc. All rights reserved.
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Mulligan, 2006). Arsenic is toxic to humans and other animals, causing neurological disorders, liver and kidney diseases, cancer, gastrointestinal disturbances and other health problems (Subhani et al., 2015). Aquatic primary producers such as phytoplankton, periphyton, bryophytes, macrophytes (Culioli et al., 2009) and macroinvertebrates accumulate arsenic from water and sediment (Awrahman et al., 2015; Magalhães et al., 2015; Rahman et al., 2012), causing reduced survival (Canivet et al., 2001; Gaete and Chávez, 2008), growth (Marsden and Rainbow, 2004) and reproduction (Mogren et al., 2012). In terms of community structure, high concentrations of trace elements, including arsenic, reduce the abundance and richness of macroinvertebrates (Doi et al., 2007; Mayer-Pinto et al., 2010; Qu et al., 2010; Rhea et al., 2006) and potentially impair their functional role in the ecosystem. Understanding the arsenic pathway in biological systems is therefore important for environmental management and restoration. We used field and laboratory approaches to investigate (1) at community level, the effects of arsenic pollution on macroinvertebrate density, richness, and community composition; (2) routes of arsenic incorporation into consumer bodies; and (3) arsenic mobility in food webs, using the stable-isotope approach, and its transfer to the terrestrial environment. We hypothesized that (1) because of its toxicity, arsenic will cause decreases in invertebrate density and richness; (2) at sub-lethal levels, consumers will incorporate arsenic both from contaminated food and directly from water contact; and (3) because of contaminated food ingestion, arsenic concentration will increase through food webs, and eventually arsenic will be transferred to the terrestrial environment with the emergence of aquatic-insect consumers.
Fig. 1. Flow-chart of the experimental approach.
Tachet et al. (1987), except dipterans which were identified to family level. 2.3. Relationship between concentrations of arsenic in stream water and biota We collected water samples, biofilm and two invertebrate taxa, a scraper (Heptagenia sp.) and a collector (Hydropsyche sp.), in four stretches along the impacted stream (sites A to D; 5.3 km apart) to investigate the relationship between arsenic concentrations in the water and in organism bodies (Fig. 1). We used these two trophic groups to understand the incorporation of arsenic into food chains from biofilm and fine particulate organic matter. We did not use shredders because their density at the stream sites was low and not enough specimens were available to make similar measurements. Along the four stretches of the polluted stream, arsenic concentrations in water decreased from 40 to 2.5 μg L–1; at the reference site, arsenic concentrations were 0.3 μg L–1. Biofilm was collected by delimiting a 72 cm2 area on submerged stones (using a flask cap) and scraping the area with a toothbrush into a 30 mL plastic flask. The samples were taken to the laboratory in a cooler box and the content was passed through a pre-weighed filter (Whatman 0.45 µm GFC), dried (60 °C/48 h) and weighed. The filters were then digested in a nitric-acid and hydrogen-peroxide solution (4:1) to determine the arsenic (see below). Arsenic concentration was determined as described in 2.7.
2. Material and methods 2.1. Study area We sampled two streams; a stream receiving water from an abandoned mine (Escádia Grande, Portugal; 'impacted stream'; 4 sites, A to D; 40°04′02″N, 008°06′22″W – to 40°02′11″N, 008°05′45″W) and a 'reference stream' located in the same area (site E), but not affected by arsenic (40°03′56″N, 008°06′17″W). The streams were similar in width, depth, flow, substrate and water chemistry, except for high electrical conductivity and arsenic concentration in the impacted stream and high phosphate in the reference stream (Table 1). 2.2. Arsenic effects on invertebrate assemblages in stream We evaluated the effects of arsenic pollution (mean ~40 μg L–1; 32 – 43 μg L–1 range of 8 measurements during four seasons in the previous year; site A) on aquatic invertebrates in terms of density, rarefied richness and community composition (Fig. 1). We collected 10 random Surber samples (0.1 m2; 500 µm) at each stream site. Live samples were transported to the laboratory in water inside a cooler box, sorted alive on the same day, and invertebrates identified to genus according to
2.4. Arsenic effect on growth and survival of a selected consumer: Sericostoma vittatum
Table 1 Chemical and physical variables of the arsenic-impacted and reference streams (mean ± SE, n =5; t-test and probabilistic p-value). Lousã Mountain, Portugal (March-April 2015). NS: non-significant. Limnological variables
Reference
Impacted
t
p
Water temperature (°C) Conductivity (μS cm–1) pH Dissolved oxygen (mg L–1) O2 saturation (%) Nitrate (μg L–1) Phosphate (μg L–1) Sulfate (μg L–1) Arsenic (μg L–1)
12 ± 0 29 ± 0 6.7 ± 0.0 11 ± 0 115 ± 3 321 ± 188 53 ± 4 19 ± 4 0.3 ± 0.2
12 ± 1 36 ± 1 6.7 ± 0.0 10 ± 0 110 ± 3 366 ± 16 22 ± 2 31 ± 21 39.2 ± 11.3
0.1 11.6 0.1 1.0 1.6 0.2 7.9 0.6 10.8
NS < 0.001 NS NS NS NS < 0.001 NS < 0.001
We used the shredder Sericostoma vittatum Rambur, 1842 (Sericostomatidae, Trichoptera) as a model organism to assess the effects of arsenic concentration on growth in a laboratory experiment (Fig. 1). S. vittatum is a common North Iberian endemic species, easily maintained and previously used in laboratory experimental studies (Feio and Graça, 2000; Graça et al., 2001). We sampled larvae from a nearby stream (Candal; 40°04′44″N, 008°12′10″W). We measured the capsule width under a binocular microscope to obtain dry mass values using an equation obtained from a set of 24 specimens which were measured to the nearest 0.1 mm, oven dried (60 °C for 72 h) and weighed to the nearest 0.01 g: mass (mg)=(0.0017* size in mm)+0.0589 (R2=0.80, n=24). Specimens were placed individually in 250 mL cups with 100 mL of 133
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digestion. In the case of spiders and invertebrates we used only one specimen per sample. However, due to the small larval sizes, in some cases we pooled 3–5 larvae to make a sample. Specimens were digested in a microwave oven (Anton Paar, Multiwave Pro) in a HNO3 (65%)– H2O2 (30%) mixture (4:1) in closed Teflon vessels at 15 atm pressure and 160 °C/30 min (Favas et al., 2012). The volume was brought to 25 mL with ultra-pure water in a calibrated flask. Water samples (analyzed 4 times during the experimental period) were analyzed directly after acidification with 65% HNO3 (3 mL L–1). Biological and water samples were loaded into an Atomic Absorption Spectrophotometer (AAS, SOLAAR M Series equipment from Thermo– Unicam) with Graphite Furnace. For the arsenic standard, we used Polish certified reference material, Virginia Tobacco Leaves (CTA-VTL2). The recovery value was 94% (using certified reference tissues) and the limit of detection was 0.5 μg L–1.
reference stream water. We added ~30 g of stream sediments ( < 0.25 mm diameter; previously incinerated at 550 °C for 8 h and washed in distilled water) to cover the cup bottom. Food was provided in the form of alder leaf discs (Alnus glutinosa L.; 8 mm diameter; n=3) previously conditioned for 25 days in the polluted or in the reference stream in 0.5 mm-mesh bags. The specimens were allocated to one of two treatments (1) water and leaves from the reference stream (n =21) and (2) water from the reference stream and leaves from the impacted stream (n=21). The experiment was carried out at 18.0 ± 0.5 °C, with a 12 L:12D photoperiod; during the experiment dissolved oxygen was always > 86% saturation. After 15 days, we removed the larvae, measured their capsule width and the specimens were dried (60 °C/ 72 h). The arsenic concentration accumulated by the larvae was determined as described in 2.7. 2.5. Source of arsenic for Sericostoma vittatum
2.8. Stable isotope analysis
To determine the routes of arsenic entrance into stream invertebrates and the effect of arsenic on feeding rates (Fig. 1), we placedthe S. vittatum specimens individually in cups with sand, water and leaf discs as indicated above. However, in this case we used four treatments with 15 larvae each: (1) leaves and water from the reference stream (RLRW); (2) leaves conditioned in the reference stream and water from the impacted stream (RL-IW); (3) leaves conditioned in the impacted stream and water from the reference stream (IL-RW); and (4) leaves and water from the impacted stream (IL-IW). To measure feeding rates, leaf discs were cut from leaves in pairs, with one pair exposed to the larvae and the other pair protected from consumption by placing it in a 7×7 cm, 0.5 mm-mesh size bag, hanging from the rim of the cup. The experiment lasted for 10 days, with new leaf discs added when the previous discs were consumed. At the end, the remaining discs were oven-dried (60 °C/72 h) and weighed. Consumption was calculated as the difference between the discs exposed to the consumer and the discs protected from feeding (in litter bags), assuming that disc pairs had the same initial mass. The animals were weighed and the arsenic concentrations in the food and in the specimens were measured (see Section 2.7).
For stable isotope analysis, oven-dried samples were ground to fine powder in a mixer (Retsch MM400), weighed on a microbalance (Mettler Toledo UMX2) to the nearest 0.3 mg (animals) or 0.5 mg (biofilm, bryophytes and macrophytes) and loaded into tin capsules. The carbon and nitrogen isotopes (δ13C/δ12C and δ15N/δ14N, respectively) were measured by elemental analyzer isotope ratio mass spectrometry using a Flash EA 1112 Series elemental analyzer coupled on-line via a Finnigan ConFlo II Interface to a Thermo Delta VS Mass Spectrometer. 2.9. Data analysis To assess the effect of arsenic on aquatic invertebrate communities, we compared density and rarefied richness between the reference and impacted streams by t-test. To assess differences in taxonomic composition, abundance values were (log [x+1]) transformed and subjected to a Permutational Multivariate Analysis of Variance (PerMANOVA) (Anderson, 2001). The relationship between the arsenic concentration in water along the impacted-stream gradient and the arsenic concentration in the biota (biofilm, scrapers and collectors) was investigated by simple linear regression analysis on (log[x+1]) transformed abundance data. S. vittatum growth was determined as the difference between the final and initial mass, divided by the elapsed time in days and initial mass. Growth rates in arsenic-exposed and unexposed larvae were compared by t-test. Consumption rates and arsenic accumulation in the laboratory feeding experiment were compared among the four treatments by a two-way ANOVA, with Tukey a posteriori test. Finally, we used two-way ANOVA to compare arsenic concentrations in aquatic organisms at different trophic levels. The same analysis was done for terrestrial organisms (adult insects and spiders). We use a linear regression analysis to evaluate the mobility of arsenic across trophic levels. All analyses were performed in the R environment.
2.6. Arsenic trophic transfer across trophic levels We investigated whether (1) arsenic increases through food webs, (2) invertebrates leaving the water transport arsenic to the land, and (3) arsenic reaches consumers on land (Fig. 1). We collected biofilm (scraping stones, as indicated above), bryophytes (Brachythecium rivulare Bruch & WP Schimper), an aquatic macrophyte (Oenanthe crocata L.) and aquatic invertebrates with a kick-net (500 µm) from the reference and impacted streams. Invertebrates were classified as scrapers, collectors, filter-feeders, shredders or predators (Tachet et al., 1987; Supplementary material). To assess if arsenic is transferred to land, we collected adult insects in emergence traps (area: 1 m2) at both sites. The emergence traps remained in the streams for 15 days (April 1 – 15, 2015), with inspections every 3 days to retrieve the emerged invertebrates. We also collected spiders (terrestrial predators) with tweezers from grasses and shrubs leaning over the stream; only specimens of Pisaura spp. (Pisauridae) were retrieved from both sites. The specimens were transported in a cooler box to the laboratory, identified, oven-dried and weighed. Specimens were randomly allocated to two groups. Some specimens were used to determine arsenic concentrations (Section 2.7) while others were analyzed for nitrogen isotopes (δ14N and δ15N) and to assess their trophic position (Section 2.8).
3. Results 3.1. Arsenic effects on invertebrate assemblages The invertebrate density and rarefied richness were higher in the reference than in the impacted stream (t=2.1, df=18, p=0.04; t=5.9, df=18, p < 0.001, respectively). The invertebrate density in the reference stream was 123 ± 21 organisms m–2 (mean ± SE) and in the impacted stream was 65 ± 13 organisms m–2. The invertebrate community composition also differed between the streams (PerMANOVA, F(1,18)=2.9, p=0.02). The mayflies Ephemerella spp., Baetis spp., and Heptagenia spp. were the most common taxa in both streams.
2.7. Arsenic determination The biological material was oven-dried (60 °C/72 h) and portions of 0.3 – 0.5 g of biofilm, bryophytes and macrophytes were prepared for 134
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Fig. 2. Relationship between arsenic concentration in the water and in the biofilm, scrapers and collectors in three stretches of a stream impacted by arsenic contamination.
3.2. Relationship between concentrations of arsenic in stream water and biota The arsenic concentration in the impacted stream water decreased from Site A (impacted stream; ca. 40 μg L–1) toward downstream (Site D, 5.3 km distant; ca. 2.5 μg L–1). Arsenic concentration was higher in biofilm than in scrapers and collectors, and in all cases decreased linearly with the distance from the most impacted stream site (biofilm: F(1,11)=3.6, p=0.008, β=9.57; scrapers: F(1,21)=6.7, p=0.01, β=0.72; collectors: F(1,14)=10.6, p=0.005, β=0.74; Fig. 2).
Fig. 3. Consumption rates and arsenic accumulation by Sericostoma vittatum larvae exposed to water (W) from the reference (R) and impacted (I) stream, and fed with leaves (L) conditioned at the reference (R) and the impacted (I) sites (mean ± SE, n=5).
3.3. Arsenic effect on growth and survival of Sericostoma vittatum
Table 2 Two-way ANOVA of arsenic accumulated in different trophic levels and streams by aquatic and terrestrial organisms. NS: non-significant.
Arsenic concentration in leaves conditioned in the impacted stream was higher than in leaves conditioned in the reference stream (151.6 ± 29.9 µg g–1 vs. 1.2 ± 0.3 µg g–1; mean ± SE; t=–5.0, df =6, p=0.01). Survival of S. vittatum was 93% for larvae fed reference leaves and 86% for larvae fed arsenic-contaminated leaves. Specimens fed contaminated leaves accumulated 20-fold more arsenic than larvae fed leaves from the reference stream (26.5 ± 9.3 μg g–1 vs. 1.3 ± 0.3 μg g–1; t=– 6.5, df =7, p=0.003). However, growth rates were higher in larvae fed contaminated than uncontaminated leaves (0.16 ± 0.04% day–1 vs. 0.05 ± 0.01% day–1; t=–2.7, df =23, p=0.01). 3.4. Sources of arsenic for Sericostoma vittatum Consumption rates of S. vittatum under laboratory conditions were statistically similar across treatments, and ranged from 23 to 30 mg leaves mg larvae–1 day–1 (Fig. 3). Consumption was similar for specimens exposed to the same water (ANOVA: F(1,56)=2.7, p=0.10) and leaves (ANOVA: F(1,56)=0.2, p=0.89). All specimens fed contaminated leaf discs accumulated arsenic, independently of the origin of the water (ANOVA: F(1,16)=23.7, p < 0.001), and accumulated 22-fold more arsenic than larvae fed reference leaf discs (Fig. 3).
df
SS
MS
F-value
p
Aquatic organisms Trophic levels Streams Trophic levels: streams Residuals
7 1 6 70
6.2 27.1 10.0 9.5
0.8 27.1 1.6 0.1
6.4 198.9 12.2
< 0.001 < 0.001 < 0.001
Terrestrial organisms Trophic levels Streams Trophic levels: streams Residuals
1 1 1 22
0.2 0.1 0.1 4.5
0.2 0.1 0.1 0.2
1.2 0.4 0.6
NS NS NS
scrapers (40 ± 11 μg g–1), and filter-feeders (30 ± 12 μg g–1; Fig. 4). Adult aquatic insects collected in the emergence traps in the reference and impacted streams did not differ in their arsenic concentrations (Table 2). These adult insects had less arsenic (9 ± 1 μg g–1) than aquatic stages of the same taxa. Spiders sampled in the impacted and reference streams did not differ in their arsenic concentrations (25 ± 7 μg g–1) (Fig. 4; Table 2). The results of the stable-isotope analysis were consistent with the trophic functional feeding analysis, revealing a decrease in arsenic concentration along the food chain (Log As=−0.163*δ15N+2.81, R2=0.19). Arsenic concentrations in the body of organisms and δ15N were negatively correlated (r=−0.51; Fig. 5).
3.5. Arsenic transfer across trophic levels Arsenic concentration in aquatic biota was higher in specimens sampled in the impacted than in the reference stream, and this difference was consistent across trophic levels (Table 2). Among primary producers, bryophytes had the highest accumulation (1,760 ± 485 μg g–1; mean ± SE), followed by the biofilm (449 ± 119 μg g–1) and macrophytes (29 ± 11 μg g–1; Fig. 4). Among aquatic invertebrates, arsenic concentration was highest in shredders (313 ± 139 μg g–1), followed by collectors (140 ± 58 μg g–1), predators (70 ± 20 μg g–1),
4. Discussion We investigated arsenic bioaccumulation by stream biota, and its 135
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4.2. Arsenic levels in the body reflect environmental levels Within the ranges of arsenic concentrations tested, stream biota accumulated arsenic proportionally to the arsenic concentration in the water, as reported by other investigators (Wang and Mulligan, 2006) and for other metals (e.g. Beghelli et al., 2016). Bryophytes and biofilm accumulated the most arsenic in our study, which is consistent with previous reports (Barral-Fraga et al., 2016) and supports the suggestions that (1) these organisms can be used as bioindicators of arsenic contamination (e.g. Favas et al., 2012; Varun et al., 2012) and (2) the ingestion of contaminated biofilm and other plant material is a possible route of arsenic flow into food webs. 4.3. Leaf-litter consumers acquire arsenic mainly by food ingestion Consumers can accumulate metals and trace elements by two routes, adsorption on the body surface or ingestion of food. Adsorption from the water is considered a main source of accumulation for invertebrates in metal-impacted streams (Rahman et al., 2012; Rainbow et al., 2007). However, Zubrod et al. (2015) found that most of the copper accumulated by Gammarus was obtained from food, and Schaller et al. (2009) found that Gammarus pulex accumulates uranium through both the body surface and food. We found that arsenic incorporation by the shredder S. vittatum occurred almost exclusively by food ingestion. Given the high capacity of biofilms to adsorb arsenic and the large surface: volume ratio of fine particulate organic matter (FPOM; a food source for many stream invertebrates), which facilitates arsenic adsorption, it is likely that other shredders, scrapers and FPOM consumers could also acquire arsenic from food. Larvae fed arsenic-contaminated leaf discs accumulated about 20-fold more arsenic than the larvae fed uncontaminated leaf discs (up to ~35 µg g−1), which is above the range of arsenic accumulation levels reported for trichopterans (0.002–0.18 µg g–1) by Wang and Mulligan (2006). If these laboratory findings can be extrapolated to environmental conditions, our study showed that arsenic contamination with levels as low as 40 µg L–1 can be magnified by detritus consumers. An unexpected finding was the faster growth of specimens fed arsenic-contaminated leaves, compared to those fed uncontaminated leaves. A reduction in growth with increased metal concentration has been reported for many consumers, including snails exposed to copper (Das and Khangarot, 2011; Mountain, 2015; Ng et al., 2011) and Heptagenia and Chironomus exposed to copper and cadmium (Milani et al., 2003). Increased growth of Cinygmula (Ephemeroptera) nymphs in a stream with moderate zinc pollution compared to reference streams was also reported by Carlisle and Clements (2005), although the authors explained the difference as differences in population life histories. The rapid growth observed for specimens fed arsenic-contaminated leaves could be an indirect hormesis effect on fungi, stimulating growth and resulting in increased maceration and better quality of litter, allowing shredders to grow rapidly (e.g. Calabrese and Baldwin, 2001). This possibility needs further investigation.
Fig. 4. Arsenic concentration (mean ± SE) in animals grouped in trophic levels, emergent adult insects and spiders in the reference and impacted streams. Biof: Biofilm; Bry: Bryophytes; Macr: Macrophytes; Scr: Scrapers; Col: Collectors; Fil: Filter-feeders; Shr: Shredders; Pred: Aquatic predators; Adult: Adult insects.
Fig. 5. Relationship between the arsenic concentrations and δ15N in different invertebrate trophic groups.
mobility in aquatic food chains and transfer to the terrestrial environment. We found that (1) low levels of arsenic caused changes in the structure and composition of the aquatic invertebrate assemblages studied; (2) arsenic accumulation by aquatic biota was related to arsenic in the environment; (3) leaf-litter consumers acquired arsenic mainly by food ingestion; and (4) there was no biomagnification of arsenic across trophic levels, and little arsenic was transported to the terrestrial environments by insect emergence. 4.1. Arsenic effects on aquatic invertebrate assemblages As hypothesized, at the site with high arsenic concentration, invertebrate density and richness were lower and the community composition differed from the reference site, an effect previously reported for arsenic (e.g. Chaffin et al., 2005) and for other metals (Iwasaki and Ormerod, 2012; Quinn et al., 2003). An important point of our study was that the effect was observed at a relatively low arsenic concentration (~40 µg L–1). This level was below the “no-effect concentration” (NOEC) reported in other aquatic invertebrates (~44 – 16,430 µg L–1) (Brix et al., 2003; Tišler and Zagorc-Končan, 2002) and is below the 150 µg L–1 US EPA (1999) levels for protection of aquatic life. Although arsenic toxicity depends on its species (Magalhães et al., 2015) and on the sensitivity of the taxa, our main conclusion regarding toxicity was that relatively low levels of arsenic were able to affect the aquatic biota.
4.4. Arsenic is not dispersed to terrestrial environments by insect emergence Our results suggest that arsenic accumulated by primary consumers is not magnified through the food webs, nor is it biologically dispersed from water to land. These findings agree with other studies (Juncos et al., 2016; Rahman et al., 2012; Watanabe et al., 2008). The decrease in arsenic concentration along trophic levels may be the result of efficiency in eliminating the contaminant (Watanabe et al., 2008). For instance, Mogren et al. (2013) found that Chironomus riparius excretes arsenic to the exoskeleton between the larval and pupal stages, while Schaller et al. (2015) observed that Gammarus pulex accumulates arsenic in the cuticle. If arthropods deposit arsenic in the exoskeleton, subsequent ecdyses will eliminate the arsenic from food webs. If the 136
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553–568. http://dx.doi.org/10.1899/04-058.1. Cui, B., Zhang, Q., Zhang, K., Liu, X., Zhang, H., 2011. Analyzing trophic transfer of heavy metals for food webs in the newly-formed wetlands of the Yellow River Delta, China. Environ. Pollut. 159, 1297–1306. http://dx.doi.org/10.1016/j.envpol.2011.01.024. Culioli, J.L., Fouquoire, A., Calendini, S., Mori, C., Orsini, A., 2009. Trophic transfer of arsenic and antimony in a freshwater ecosystem: a field study. Aquat. Toxicol. 94, 286–293. http://dx.doi.org/10.1016/j.aquatox.2009.07.016. Das, S., Khangarot, B.S., 2011. Bioaccumulation of copper and toxic effects on feeding, growth, fecundity and development of pond snail Lymnaea luteola L. J. Hazard. Mater. 185, 295–305. http://dx.doi.org/10.1016/j.jhazmat.2010.09.033. Doi, H., Takagi, A., Kikuchi, E., 2007. Stream Macroinvertebrate Community Affected by Point-Source Metal Pollution. Int. Rev. Hydrobiol. 92, 258–266. http://dx.doi.org/ 10.1002/iroh.200610923. Favas, P.J., Pratas, J., Prasad, M.N., 2012. Accumulation of arsenic by aquatic plants in large-scale field conditions: opportunities for phytoremediation and bioindication. Sci. Total Environ. 433, 390–397. http://dx.doi.org/10.1016/ j.scitotenv.2012.06.091. Feio, M.J., Graça, M.A.S., 2000. Food consumption by the larvae of Sericostoma vittatum (Trichoptera), an endemic species from the Iberian Peninsula. Hydrobiologia 439, 7–11. http://dx.doi.org/10.1023/A:1004189316952. Gaete, H., Chávez, C., 2008. Evaluación de la toxicidad de mezclas binarias de cobre, zinc y arsénico sobre Daphnia obtusa (Kurz, 1874) (Cladocera, Crustacea). Limnetica 27, 1–10. Gall, J.E., Boyd, R.S., Rajakaruna, N., 2015. Transfer of heavy metals through terrestrial food webs: a review. Environ. Monit. Assess. 187, 201–222. http://dx.doi.org/ 10.1007/s10661-015-4436-3. Ghaeni, M., Pour, N.A., Hosseini, M., 2015. Bioaccumulation of polychlorinated biphenyl (PCB), polycyclic aromatic hydrocarbon (PAH), mercury, methyl mercury, and arsenic in blue crab Portunus segnis from Persian Gulf. Environ. Monit. Assess. 187, 253–262. http://dx.doi.org/10.1007/s10661-015-4459-9. Graça, M.A.S., Cressa, C., Gessner, M.O., Feio, M.J., Callies, K.A., Barrios, C., 2001. Food quality, feeding preferences, survival and growth of shredders from temperate and tropical streams. Freshwat. Biol. 46, 947–957. http://dx.doi.org/10.1046/j.13652427.2001.00729.x. Ikemoto, T., Tu, N.P., Okuda, N., Iwata, A., Omori, K., Tanabe, S., Tuyen, B.C., Takeuchi, I., 2008. Biomagnification of trace elements in the aquatic food web in the Mekong Delta, South Vietnam using stable carbon and nitrogen isotope analysis. Arch. Environ. Contam. Toxicol. 54, 504–515. http://dx.doi.org/10.1007/s00244-0079058-5. Iwasaki, Y., Ormerod, S.J., 2012. Estimating safe concentrations of trace metals from inter-continental field data on river macroinvertebrates. Environ. Pollut. 166, 182–186. http://dx.doi.org/10.1016/j.envpol.2012.03.028. Juncos, R., Arcagni, M., Rizzo, A., Campbell, L., Arribere, M., Guevara, S.R., 2016. Natural origin arsenic in aquatic organisms from a deep oligotrophic lake under the influence of volcanic eruptions. Chemosphere 144, 2277–2289. http://dx.doi.org/ 10.1016/j.chemosphere.2015.10.092. Kehrig, H.A., Seixas, T.G., Malm, O., Di Beneditto, A.P., Rezende, C.E., 2013. Mercury and selenium biomagnification in a Brazilian coastal food web using nitrogen stable isotope analysis: a case study in an area under the influence of the Paraiba do Sul River plume. Mar. Pollut. Bull. 75, 283–290. http://dx.doi.org/10.1016/ j.marpolbul.2013.06.046. Liu, C.P., Luo, C.L., Gao, Y., Li, F.B., Lin, L.W., Wu, C.A., Li, X.D., 2010. Arsenic contamination and potential health risk implications at an abandoned tungsten mine, southern China. Environ. Pollut. 158, 820–826. http://dx.doi.org/10.1016/ j.envpol.2009.09.029. Magalhães, D.P., Marques, M.R.C., Baptista, D.,F., Buss, D.F., 2015. Metal bioavailability and toxicity in freshwaters. Environ. Chem. Lett. 13, 69–87. http://dx.doi.org/ 10.1007/s10311-015-0491-9. Marsden, I.D., Rainbow, P.S., 2004. Does the accumulation of trace metals in crustaceans affect their ecology — the amphipod example? J. Exp. Mar. Biol. Ecol. 300, 373–408. http://dx.doi.org/10.1016/j.jembe.2003.12.009. Mayer-Pinto, M., Underwood, A.J., Tolhurst, T., Coleman, R.A., 2010. Effects of metals on aquatic assemblages: what do we really know? J. Exp. Mar. Biol. Ecol. 391, 1–9. http://dx.doi.org/10.1016/j.jembe.2010.06.013. Milani, D., Reynoldson, T.B., Borgmann, U., Kolasa, J., 2003. The relative sensitivity of four benthic invertebrates to metals in spiked-sediment exposures and application to contaminated field sediment. Environ. Toxicol. Chem. 22, 845–854. http:// dx.doi.org/10.1002/etc.5620220424. Mogren, C.L., von Kiparski, G.R., Parker, D.R., Trumble, J.T., 2012. Survival, reproduction, and arsenic body burdens in Chironomus riparius exposed to arsenate and phosphate. Sci. Total Environ. 425, 60–65. http://dx.doi.org/10.1016/ j.scitotenv.2012.03.009 (DOI: 10.1016/j.scitotenv.2012.03.009). Mogren, C.L., Walton, W.E., Parker, D.R., Trumble, J.T., 2013. Trophic transfer of arsenic from an aquatic insect to terrestrial insect predators. PLoS One 8, e67817. http:// dx.doi.org/10.1371/journal.pone.0067817. Mountain, R.I.E.W.G., 2015. Elevation-dependent warming in mountain regions of the world. Nat. Clim. Change 5, 424–430. http://dx.doi.org/10.1038/NCLIMATE2563. Ng, T.Y.T., Pais, N.M., Wood, C.M., 2011. Mechanisms of waterborne Cu toxicity to the pond snail Lymnaea stagnalis: physiology and Cu bioavailability. Ecotoxicol. Environ. Saf. 74, 1471–1479. http://dx.doi.org/10.1016/j.ecoenv.2011.03.010. Painter, K.J., Janz, D.M., Jardine, T.D., 2016. Bioaccumulation of mercury in invertebrate food webs of Canadian Rocky Mountain streams. Freshw. Sci. 35, 1248–1262. http:// dx.doi.org/10.1086/688897. Qu, X., Wu, N., Tang, T., Cai, Q., Park, Y.S., 2010. Effects of heavy metals on benthic macroinvertebrate communities in high mountain streams. Ann. Limnol. - Int. J. Limnol. 46, 291–302. http://dx.doi.org/10.1051/limn/2010027.
process is efficient, then arsenic decreases through food webs, and adults may transport little arsenic when leaving the aquatic environment. At the arsenic levels observed in the stream (40 μg L–1), the mechanism for arsenic elimination was efficient, since we observed no differences in arsenic concentration in spiders living over the arsenic-impacted and reference streams. If spiders are good models of terrestrial consumers, we can conclude that there is no biodispersal of arsenic from water to land. 5. Conclusions At very low concentrations (~40 μg L–1), arsenic affected invertebrate assemblages, with changes in composition and taxa richness. Arsenic in streams is adsorbed by producers and by dead organic matter proportionally to its concentration in the water, and in this way enters food webs. Once incorporated into the producers, it can pass to consumers, but it is not biomagnified. Apparently, invertebrates have efficient mechanisms for arsenic elimination. It is possible for arsenic to be locked in the exoskeleton and eliminated during the ecdysis. Therefore, aquatic organisms do not cause the biodispersal of arsenic to terrestrial systems. Acknowledgements The authors thank Igor Moraes, Joana Sotaia Campos and Cristina Docal for their help in the field and laboratory work. Seena Sahadevan critically reviewed the manuscript. Rodrigo Fornel for the graphic abstract edition. LUH received support from the Programa Ciências Sem Fronteiras/CNPq/Brazil (Proc. #250080/2013-3). This research was supported by the Marine and Environmental Sciences Centre (MARE) through the program UID/MAR/04292/2013. Appendix A. Supporting information Supplementary data associated with this article can be found in the online version at doi:10.1016/j.ecoenv.2017.01.035. References Alonso, D.L., Latorre, S., Castillo, E., Brandão, P.F.B., 2014. Environmental occurrence of arsenic in Colombia: a review. Environ. Pollut. 186, 272–281. http://dx.doi.org/ 10.1016/j.envpol.2013.12.009. Ancion, P.Y., Lear, G., Lewis, G.D., 2010. Three common metal contaminants of urban runoff (Zn, Cu & Pb) accumulate in freshwater biofilm and modify embedded bacterial communities. Environ. Pollut. 158, 2738–2745. http://dx.doi.org/10.1016/ j.envpol.2010.04.013. Anderson, M.J., 2001. A new method for non-parametric multivariate analysis of variance. Austral Ecol. 26, 32–46. http://dx.doi.org/10.1111/j.14429993.2001.01070.pp.x. Awrahman, Z.A., Rainbow, P.S., Smith, B.D., Khan, F.R., Bury, N.R., Fialkowski, W., 2015. Bioaccumulation of arsenic and silver by the caddisfly larvae Hydropsyche siltalai and H. pellucidula: a biodynamic modeling approach. Aquat. Toxicol. 161, 196–207. http://dx.doi.org/10.1016/j.aquatox.2015.01.004. Barral-Fraga, L., Morin, S., Rovira, M.D.M., Urrea, G., Magellan, K., Guasch, H., 2016. Short-term arsenic exposure reduces diatom cell size in biofilm communities. Environ. Sci. Pollut. Res. 23, 4257–4270. http://dx.doi.org/10.1007/s11356-0154894-8. Beghelli, F.G.S., Pompêo, M.L.M., Rosa, A.H., Moschini-Carlos, V., 2016. Effects of copper in sediments on benthic macroinvertebrate communities in tropical reservoirs. Limnetica 35, 103–116. Brix, K.V., Cardwell, R.D., Adams, W.J., 2003. Chronic toxicity of arsenic to the Great salt Lake brine shrimp, Artemia franciscana. Ecotoxicol. Environ. Saf. 54, 169–175. Calabrese, E.J., Baldwin, L.A., 2001. Hormesis: a generalizable and unifying hypothesis. Crit. Rev. Toxicol. 31, 353–424. http://dx.doi.org/10.1080/20014091111730. Canivet, V., Chamnon, P., Gilbert, J., 2001. Toxicity and bioaccumulation of Arsenic and Chromium in epigean and hypogean freshwater macroinvertebrates. Arch. Environ. Contam Toxicol. 40, 345–354. http://dx.doi.org/10.1007/s002440010182. Carlisle, D.M., Clements, W.H., 2005. Leaf litter breakdown, microbial respiration and shredder production in metal-polluted streams. Freshw. Biol. 50, 380–390. http:// dx.doi.org/10.1111/j.1365-2427.2004.01323.x. Chaffin, J.L., Valett, H.M., Webster, J.R., 2005. Influence of elevated As on leaf breakdown in an Appalachian headwater stream. J. N. Am. Benthol. Soc. 24,
137
Ecotoxicology and Environmental Safety 139 (2017) 132–138
L.U. Hepp et al.
des eaux douces (Systématique élémentaire et aperçu écologique). Université Lyon, Lyon. Tišler, T., Zagorc-Končan, J., 2002. Acute and chronic toxicity of arsenic to some aquatic organisms. Bull. Environ. Contam. Toxicol. 69, 421–429. http://dx.doi.org/10.1007/ s00128-002-0079-5. Varun, M., D’Souza, R., Pratas, J., Paul, M.S., 2012. Metal contamination of soils and plants associated with the glass industry in North Central India: prospects of phytoremediation. Environ. Sci. Pollut. Res. 19, 269–281. http://dx.doi.org/ 10.1007/s11356-011-0530-4. Vithanage, M., Dabrowska, B.B., Mukherjee, A.B., Sandhi, A., Bhattacharya, P., 2011. Arsenic uptake by plants and possible phytoremediation applications: a brief overview. Environ. Chem. Lett. 10, 217–224. http://dx.doi.org/10.1007/s10311011-0349-8. Wang, S., Mulligan, C.N., 2006. Occurrence of arsenic contamination in Canada: sources, behavior and distribution. Sci. Total Environ. 366, 701–721. http://dx.doi.org/ 10.1016/j.scitotenv.2005.09.005. Wang, X., Peng, B., Tan, C., Ma, L., Rathinasabapathi, B., 2015. Recent advances in arsenic bioavailability, transport, and speciation in rice. Environ. Sci. Pollut. Res. 22, 5742–5750. http://dx.doi.org/10.1007/s11356-014-4065-3. Watanabe, K., Monaghan, M.T., Takemond, Y., Omura, T., 2008. Biodilution of heavy metals in a stream macroinvertebrate food web: evidence from stable isotope analysis. Sci. Total Environ. 394, 57–67. http://dx.doi.org/10.1016/ j.scitotenv.2008.01.006. Zubrod, J.P., Englert, D., Rosenfeldt, R.R., Wolfram, J., Lüderwald, S., Wallace, D., Schnetzer, N., Schulz, R., Bundschuh, M., 2015. The relative importance of dietrelated and waterborne effects of copper for a leaf-shredding invertebrate. Environ. Pollut. 205, 16–22.
Quinn, M.R., Feng, X., Folt, C.L., Chamberlain, C.P., 2003. Analyzing trophic transfer of metals in stream food webs using nitrogen isotopes. Sci. Total Environ. 317, 73–89. http://dx.doi.org/10.1016/s0048-9697(02)00615-0. Rahman, M.A., Hasegawa, H., Lim, R.P., 2012. Bioaccumulation, biotransformation and trophic transfer of arsenic in the aquatic food chain. Environ. Res. 116, 118–135. http://dx.doi.org/10.1016/j.envres.2012.03.014. Rainbow, P.S., Amiard, J.C., Amiard-Triquet, C., Cheung, M., Zhang, L., Zhong, H., Wang, W., 2007. Trophic transfer of trace metals: subcellular compartmentalization in bivalve prey, assimilation by a gastropod predator and in vitro digestion simulations. Mar. Ecol. Prog. Ser. 348, 125–138. http://dx.doi.org/10.1016/ j.envint.2006.05.007. Rhea, D.T., Harper, D.D., Farag, A.M., Brumbaugh, W.G., 2006. Biomonitoring in the Boulder River watershed, Montana, USA: metal concentrations in biofilm and macroinvertebrates, and relations with macroinvertebrate assemblage. Environ. Monit. Assess. 115, 381–393. http://dx.doi.org/10.1007/s10661-006-7086-7. Schaller, J., Brackhage, C., Dudel, E.G., 2009. Limited transfer of uranium to higher trophic levels by Gammarus pulex L. in contaminated environments. J. Environ. Monit. 11, 1629–1633. http://dx.doi.org/10.1039/b906420f. Schaller, J., Kochb, I., Caumetteb, G., Nearingb, M., Reimerb, K.J., Planer-Friedricha, B., 2015. Strategies of Gammarus pulex L. to cope with arsenic – results from speciation analyses by IC-ICP-MS and XAS micro-mapping. Sci. Total Environ. 530–531, 430–433. http://dx.doi.org/10.1016/j.scitotenv.2015.06.015. Subhani, M., Mustafa, I., Alamdar, A., Katsoyiannis, I.A., Ali, N., Huang, Q., Peng, S., Shen, H., Eqani, S.A., 2015. Arsenic levels from different land-use settings in Pakistan: bio-accumulation and estimation of potential human health risk via dust exposure. Ecotoxicol. Environ. Saf. 115, 187–194. http://dx.doi.org/10.1016/ j.ecoenv.2015.02.019. Tachet, H., Bournaud, M., Richoux, P., 1987. Introduction à l′étude des macroinvertébrés
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Arsenic in stream waters is bioaccumulated but neither biomagnified through food webs nor biodispersed to land
MARK
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Luiz U. Heppa, , João A.M.S. Pratasb, Manuel A.S. Graçac a b c
Departament of Biological Sciences, Universidade Regional Integrada do Alto Uruguai e das Missões, 99709-910 Rio Grande do Sul, Brazil MARE – Marine and Environmental Sciences Centre, Department of Earth Sciences, University of Coimbra, 3001-517 Coimbra, Portugal MARE – Marine and Environmental Sciences Centre, Department of Life Sciences, University of Coimbra, 3001-401 Coimbra, Portugal
A R T I C L E I N F O
A B S T R A C T
Keywords: Biomagnification Isotopic approach Arsenic contamination Macroinvertebrates Water quality
Human activities such as mining have contributed substantially to the increase of metals in aquatic environments worldwide. These metals are bioaccumulated by aquatic organisms and can be biomagnified along trophic webs. The dispersal of contaminants from water to land has been little investigated, even though most aquatic invertebrates in streams have aerial stages. We used field and laboratory approaches to investigate the effects of arsenic pollution on stream invertebrate assemblages, and its bioaccumulation, biomagnification and trophic transfer from aquatic to terrestrial environments by emergent insects. We conducted the study in an arsenicimpacted stream (40 μg L–1 As at the most polluted site) and a reference stream (0.3 μg L–1 As). Invertebrate abundance and richness were lowest at the most impacted site. Arsenic in biofilm and in invertebrates increased with the arsenic content in the water. The highest arsenic accumulators were bryophytes (1760 μg g–1), followed by the biofilm (449 μg g–1) and shredder invertebrates (313 μg g–1); predators had the lowest arsenic concentration. Insects emerging from water and spiders along streambanks sampled from the reference and the impacted stream did not differ in their body arsenic concentrations. In the laboratory, the shredder Sericostoma vittatum had reduced feeding rates when exposed to water from the impacted stream in comparison with the reference stream (15.6 vs. 19.0 mg leaves mg body mass–1 day–1; p < 0.05), but they grew faster in the polluted water (0.16 ± 0.04% day–1 vs. 0.05 ± 0.01% day–1, p < 0.05). S. vittatum exposed to contaminated stream water accumulated arsenic from food, not through contact with water. We concluded that although arsenic is bioaccumulated, mainly by food ingestion, it is not biomagnified through food webs and is not transported from the aquatic to terrestrial environment when insects leave the stream water.
1. Introduction As a consequence of human activities, many natural and artificial compounds are being released into the environment. Many such compounds enter the food webs and profoundly harm the environment and humans. Some pollutants are accumulated by living organisms, particularly by primary producers (Ancion et al., 2010; Varun et al., 2012; Vithanage et al., 2011). When the concentration of pollutants is higher in the organisms than in their environment, there is a possibility of biomagnification through food webs, particularly affecting higher trophic levels. If highly mobile organisms bioaccumulate pollutants, they can disperse these pollutants away from the source of contamination. Most invertebrates in low-order streams are insects with an aerial stage, and therefore capable of transporting pollutants to land (Mogren et al., 2013). We investigated the potential movement of a pollutant
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from water to land by insects, using arsenic as the model. Metals and metalloids are common in nature, but can be toxic even at low concentrations (Gall et al., 2015). Biomagnification of Cd, Zn, Hg and Se was observed in freshwaters, but the mobility of other elements through food webs remains unclear (Cui et al., 2011; Ikemoto et al., 2008; Juncos et al., 2016; Kehrig et al., 2013; Painter et al., 2016). Arsenic is a toxic trace element that occurs naturally in the environment. Arsenic in water may originate from rocks by weathering (Alonso et al., 2014; Wang and Mulligan, 2006) and volcanic activity (Juncos et al., 2016). Human activities, including industry, municipal wastes and particularly mining, are important sources of arsenic to the environment (Liu et al., 2010; Subhani et al., 2015). Fungicides, herbicides, insecticides and phosphate fertilizers are also potential sources of arsenic contamination (Ghaeni et al., 2015; Wang and Mulligan, 2006; Wang et al., 2015). All organisms can take up arsenic (as arsenite) because it is chemically similar to phosphate (Wang and
Correspondence author. E-mail address: [email protected] (L.U. Hepp).
http://dx.doi.org/10.1016/j.ecoenv.2017.01.035 Received 28 September 2016; Received in revised form 16 January 2017; Accepted 19 January 2017 0147-6513/ © 2017 Elsevier Inc. All rights reserved.
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Mulligan, 2006). Arsenic is toxic to humans and other animals, causing neurological disorders, liver and kidney diseases, cancer, gastrointestinal disturbances and other health problems (Subhani et al., 2015). Aquatic primary producers such as phytoplankton, periphyton, bryophytes, macrophytes (Culioli et al., 2009) and macroinvertebrates accumulate arsenic from water and sediment (Awrahman et al., 2015; Magalhães et al., 2015; Rahman et al., 2012), causing reduced survival (Canivet et al., 2001; Gaete and Chávez, 2008), growth (Marsden and Rainbow, 2004) and reproduction (Mogren et al., 2012). In terms of community structure, high concentrations of trace elements, including arsenic, reduce the abundance and richness of macroinvertebrates (Doi et al., 2007; Mayer-Pinto et al., 2010; Qu et al., 2010; Rhea et al., 2006) and potentially impair their functional role in the ecosystem. Understanding the arsenic pathway in biological systems is therefore important for environmental management and restoration. We used field and laboratory approaches to investigate (1) at community level, the effects of arsenic pollution on macroinvertebrate density, richness, and community composition; (2) routes of arsenic incorporation into consumer bodies; and (3) arsenic mobility in food webs, using the stable-isotope approach, and its transfer to the terrestrial environment. We hypothesized that (1) because of its toxicity, arsenic will cause decreases in invertebrate density and richness; (2) at sub-lethal levels, consumers will incorporate arsenic both from contaminated food and directly from water contact; and (3) because of contaminated food ingestion, arsenic concentration will increase through food webs, and eventually arsenic will be transferred to the terrestrial environment with the emergence of aquatic-insect consumers.
Fig. 1. Flow-chart of the experimental approach.
Tachet et al. (1987), except dipterans which were identified to family level. 2.3. Relationship between concentrations of arsenic in stream water and biota We collected water samples, biofilm and two invertebrate taxa, a scraper (Heptagenia sp.) and a collector (Hydropsyche sp.), in four stretches along the impacted stream (sites A to D; 5.3 km apart) to investigate the relationship between arsenic concentrations in the water and in organism bodies (Fig. 1). We used these two trophic groups to understand the incorporation of arsenic into food chains from biofilm and fine particulate organic matter. We did not use shredders because their density at the stream sites was low and not enough specimens were available to make similar measurements. Along the four stretches of the polluted stream, arsenic concentrations in water decreased from 40 to 2.5 μg L–1; at the reference site, arsenic concentrations were 0.3 μg L–1. Biofilm was collected by delimiting a 72 cm2 area on submerged stones (using a flask cap) and scraping the area with a toothbrush into a 30 mL plastic flask. The samples were taken to the laboratory in a cooler box and the content was passed through a pre-weighed filter (Whatman 0.45 µm GFC), dried (60 °C/48 h) and weighed. The filters were then digested in a nitric-acid and hydrogen-peroxide solution (4:1) to determine the arsenic (see below). Arsenic concentration was determined as described in 2.7.
2. Material and methods 2.1. Study area We sampled two streams; a stream receiving water from an abandoned mine (Escádia Grande, Portugal; 'impacted stream'; 4 sites, A to D; 40°04′02″N, 008°06′22″W – to 40°02′11″N, 008°05′45″W) and a 'reference stream' located in the same area (site E), but not affected by arsenic (40°03′56″N, 008°06′17″W). The streams were similar in width, depth, flow, substrate and water chemistry, except for high electrical conductivity and arsenic concentration in the impacted stream and high phosphate in the reference stream (Table 1). 2.2. Arsenic effects on invertebrate assemblages in stream We evaluated the effects of arsenic pollution (mean ~40 μg L–1; 32 – 43 μg L–1 range of 8 measurements during four seasons in the previous year; site A) on aquatic invertebrates in terms of density, rarefied richness and community composition (Fig. 1). We collected 10 random Surber samples (0.1 m2; 500 µm) at each stream site. Live samples were transported to the laboratory in water inside a cooler box, sorted alive on the same day, and invertebrates identified to genus according to
2.4. Arsenic effect on growth and survival of a selected consumer: Sericostoma vittatum
Table 1 Chemical and physical variables of the arsenic-impacted and reference streams (mean ± SE, n =5; t-test and probabilistic p-value). Lousã Mountain, Portugal (March-April 2015). NS: non-significant. Limnological variables
Reference
Impacted
t
p
Water temperature (°C) Conductivity (μS cm–1) pH Dissolved oxygen (mg L–1) O2 saturation (%) Nitrate (μg L–1) Phosphate (μg L–1) Sulfate (μg L–1) Arsenic (μg L–1)
12 ± 0 29 ± 0 6.7 ± 0.0 11 ± 0 115 ± 3 321 ± 188 53 ± 4 19 ± 4 0.3 ± 0.2
12 ± 1 36 ± 1 6.7 ± 0.0 10 ± 0 110 ± 3 366 ± 16 22 ± 2 31 ± 21 39.2 ± 11.3
0.1 11.6 0.1 1.0 1.6 0.2 7.9 0.6 10.8
NS < 0.001 NS NS NS NS < 0.001 NS < 0.001
We used the shredder Sericostoma vittatum Rambur, 1842 (Sericostomatidae, Trichoptera) as a model organism to assess the effects of arsenic concentration on growth in a laboratory experiment (Fig. 1). S. vittatum is a common North Iberian endemic species, easily maintained and previously used in laboratory experimental studies (Feio and Graça, 2000; Graça et al., 2001). We sampled larvae from a nearby stream (Candal; 40°04′44″N, 008°12′10″W). We measured the capsule width under a binocular microscope to obtain dry mass values using an equation obtained from a set of 24 specimens which were measured to the nearest 0.1 mm, oven dried (60 °C for 72 h) and weighed to the nearest 0.01 g: mass (mg)=(0.0017* size in mm)+0.0589 (R2=0.80, n=24). Specimens were placed individually in 250 mL cups with 100 mL of 133
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digestion. In the case of spiders and invertebrates we used only one specimen per sample. However, due to the small larval sizes, in some cases we pooled 3–5 larvae to make a sample. Specimens were digested in a microwave oven (Anton Paar, Multiwave Pro) in a HNO3 (65%)– H2O2 (30%) mixture (4:1) in closed Teflon vessels at 15 atm pressure and 160 °C/30 min (Favas et al., 2012). The volume was brought to 25 mL with ultra-pure water in a calibrated flask. Water samples (analyzed 4 times during the experimental period) were analyzed directly after acidification with 65% HNO3 (3 mL L–1). Biological and water samples were loaded into an Atomic Absorption Spectrophotometer (AAS, SOLAAR M Series equipment from Thermo– Unicam) with Graphite Furnace. For the arsenic standard, we used Polish certified reference material, Virginia Tobacco Leaves (CTA-VTL2). The recovery value was 94% (using certified reference tissues) and the limit of detection was 0.5 μg L–1.
reference stream water. We added ~30 g of stream sediments ( < 0.25 mm diameter; previously incinerated at 550 °C for 8 h and washed in distilled water) to cover the cup bottom. Food was provided in the form of alder leaf discs (Alnus glutinosa L.; 8 mm diameter; n=3) previously conditioned for 25 days in the polluted or in the reference stream in 0.5 mm-mesh bags. The specimens were allocated to one of two treatments (1) water and leaves from the reference stream (n =21) and (2) water from the reference stream and leaves from the impacted stream (n=21). The experiment was carried out at 18.0 ± 0.5 °C, with a 12 L:12D photoperiod; during the experiment dissolved oxygen was always > 86% saturation. After 15 days, we removed the larvae, measured their capsule width and the specimens were dried (60 °C/ 72 h). The arsenic concentration accumulated by the larvae was determined as described in 2.7. 2.5. Source of arsenic for Sericostoma vittatum
2.8. Stable isotope analysis
To determine the routes of arsenic entrance into stream invertebrates and the effect of arsenic on feeding rates (Fig. 1), we placedthe S. vittatum specimens individually in cups with sand, water and leaf discs as indicated above. However, in this case we used four treatments with 15 larvae each: (1) leaves and water from the reference stream (RLRW); (2) leaves conditioned in the reference stream and water from the impacted stream (RL-IW); (3) leaves conditioned in the impacted stream and water from the reference stream (IL-RW); and (4) leaves and water from the impacted stream (IL-IW). To measure feeding rates, leaf discs were cut from leaves in pairs, with one pair exposed to the larvae and the other pair protected from consumption by placing it in a 7×7 cm, 0.5 mm-mesh size bag, hanging from the rim of the cup. The experiment lasted for 10 days, with new leaf discs added when the previous discs were consumed. At the end, the remaining discs were oven-dried (60 °C/72 h) and weighed. Consumption was calculated as the difference between the discs exposed to the consumer and the discs protected from feeding (in litter bags), assuming that disc pairs had the same initial mass. The animals were weighed and the arsenic concentrations in the food and in the specimens were measured (see Section 2.7).
For stable isotope analysis, oven-dried samples were ground to fine powder in a mixer (Retsch MM400), weighed on a microbalance (Mettler Toledo UMX2) to the nearest 0.3 mg (animals) or 0.5 mg (biofilm, bryophytes and macrophytes) and loaded into tin capsules. The carbon and nitrogen isotopes (δ13C/δ12C and δ15N/δ14N, respectively) were measured by elemental analyzer isotope ratio mass spectrometry using a Flash EA 1112 Series elemental analyzer coupled on-line via a Finnigan ConFlo II Interface to a Thermo Delta VS Mass Spectrometer. 2.9. Data analysis To assess the effect of arsenic on aquatic invertebrate communities, we compared density and rarefied richness between the reference and impacted streams by t-test. To assess differences in taxonomic composition, abundance values were (log [x+1]) transformed and subjected to a Permutational Multivariate Analysis of Variance (PerMANOVA) (Anderson, 2001). The relationship between the arsenic concentration in water along the impacted-stream gradient and the arsenic concentration in the biota (biofilm, scrapers and collectors) was investigated by simple linear regression analysis on (log[x+1]) transformed abundance data. S. vittatum growth was determined as the difference between the final and initial mass, divided by the elapsed time in days and initial mass. Growth rates in arsenic-exposed and unexposed larvae were compared by t-test. Consumption rates and arsenic accumulation in the laboratory feeding experiment were compared among the four treatments by a two-way ANOVA, with Tukey a posteriori test. Finally, we used two-way ANOVA to compare arsenic concentrations in aquatic organisms at different trophic levels. The same analysis was done for terrestrial organisms (adult insects and spiders). We use a linear regression analysis to evaluate the mobility of arsenic across trophic levels. All analyses were performed in the R environment.
2.6. Arsenic trophic transfer across trophic levels We investigated whether (1) arsenic increases through food webs, (2) invertebrates leaving the water transport arsenic to the land, and (3) arsenic reaches consumers on land (Fig. 1). We collected biofilm (scraping stones, as indicated above), bryophytes (Brachythecium rivulare Bruch & WP Schimper), an aquatic macrophyte (Oenanthe crocata L.) and aquatic invertebrates with a kick-net (500 µm) from the reference and impacted streams. Invertebrates were classified as scrapers, collectors, filter-feeders, shredders or predators (Tachet et al., 1987; Supplementary material). To assess if arsenic is transferred to land, we collected adult insects in emergence traps (area: 1 m2) at both sites. The emergence traps remained in the streams for 15 days (April 1 – 15, 2015), with inspections every 3 days to retrieve the emerged invertebrates. We also collected spiders (terrestrial predators) with tweezers from grasses and shrubs leaning over the stream; only specimens of Pisaura spp. (Pisauridae) were retrieved from both sites. The specimens were transported in a cooler box to the laboratory, identified, oven-dried and weighed. Specimens were randomly allocated to two groups. Some specimens were used to determine arsenic concentrations (Section 2.7) while others were analyzed for nitrogen isotopes (δ14N and δ15N) and to assess their trophic position (Section 2.8).
3. Results 3.1. Arsenic effects on invertebrate assemblages The invertebrate density and rarefied richness were higher in the reference than in the impacted stream (t=2.1, df=18, p=0.04; t=5.9, df=18, p < 0.001, respectively). The invertebrate density in the reference stream was 123 ± 21 organisms m–2 (mean ± SE) and in the impacted stream was 65 ± 13 organisms m–2. The invertebrate community composition also differed between the streams (PerMANOVA, F(1,18)=2.9, p=0.02). The mayflies Ephemerella spp., Baetis spp., and Heptagenia spp. were the most common taxa in both streams.
2.7. Arsenic determination The biological material was oven-dried (60 °C/72 h) and portions of 0.3 – 0.5 g of biofilm, bryophytes and macrophytes were prepared for 134
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Fig. 2. Relationship between arsenic concentration in the water and in the biofilm, scrapers and collectors in three stretches of a stream impacted by arsenic contamination.
3.2. Relationship between concentrations of arsenic in stream water and biota The arsenic concentration in the impacted stream water decreased from Site A (impacted stream; ca. 40 μg L–1) toward downstream (Site D, 5.3 km distant; ca. 2.5 μg L–1). Arsenic concentration was higher in biofilm than in scrapers and collectors, and in all cases decreased linearly with the distance from the most impacted stream site (biofilm: F(1,11)=3.6, p=0.008, β=9.57; scrapers: F(1,21)=6.7, p=0.01, β=0.72; collectors: F(1,14)=10.6, p=0.005, β=0.74; Fig. 2).
Fig. 3. Consumption rates and arsenic accumulation by Sericostoma vittatum larvae exposed to water (W) from the reference (R) and impacted (I) stream, and fed with leaves (L) conditioned at the reference (R) and the impacted (I) sites (mean ± SE, n=5).
3.3. Arsenic effect on growth and survival of Sericostoma vittatum
Table 2 Two-way ANOVA of arsenic accumulated in different trophic levels and streams by aquatic and terrestrial organisms. NS: non-significant.
Arsenic concentration in leaves conditioned in the impacted stream was higher than in leaves conditioned in the reference stream (151.6 ± 29.9 µg g–1 vs. 1.2 ± 0.3 µg g–1; mean ± SE; t=–5.0, df =6, p=0.01). Survival of S. vittatum was 93% for larvae fed reference leaves and 86% for larvae fed arsenic-contaminated leaves. Specimens fed contaminated leaves accumulated 20-fold more arsenic than larvae fed leaves from the reference stream (26.5 ± 9.3 μg g–1 vs. 1.3 ± 0.3 μg g–1; t=– 6.5, df =7, p=0.003). However, growth rates were higher in larvae fed contaminated than uncontaminated leaves (0.16 ± 0.04% day–1 vs. 0.05 ± 0.01% day–1; t=–2.7, df =23, p=0.01). 3.4. Sources of arsenic for Sericostoma vittatum Consumption rates of S. vittatum under laboratory conditions were statistically similar across treatments, and ranged from 23 to 30 mg leaves mg larvae–1 day–1 (Fig. 3). Consumption was similar for specimens exposed to the same water (ANOVA: F(1,56)=2.7, p=0.10) and leaves (ANOVA: F(1,56)=0.2, p=0.89). All specimens fed contaminated leaf discs accumulated arsenic, independently of the origin of the water (ANOVA: F(1,16)=23.7, p < 0.001), and accumulated 22-fold more arsenic than larvae fed reference leaf discs (Fig. 3).
df
SS
MS
F-value
p
Aquatic organisms Trophic levels Streams Trophic levels: streams Residuals
7 1 6 70
6.2 27.1 10.0 9.5
0.8 27.1 1.6 0.1
6.4 198.9 12.2
< 0.001 < 0.001 < 0.001
Terrestrial organisms Trophic levels Streams Trophic levels: streams Residuals
1 1 1 22
0.2 0.1 0.1 4.5
0.2 0.1 0.1 0.2
1.2 0.4 0.6
NS NS NS
scrapers (40 ± 11 μg g–1), and filter-feeders (30 ± 12 μg g–1; Fig. 4). Adult aquatic insects collected in the emergence traps in the reference and impacted streams did not differ in their arsenic concentrations (Table 2). These adult insects had less arsenic (9 ± 1 μg g–1) than aquatic stages of the same taxa. Spiders sampled in the impacted and reference streams did not differ in their arsenic concentrations (25 ± 7 μg g–1) (Fig. 4; Table 2). The results of the stable-isotope analysis were consistent with the trophic functional feeding analysis, revealing a decrease in arsenic concentration along the food chain (Log As=−0.163*δ15N+2.81, R2=0.19). Arsenic concentrations in the body of organisms and δ15N were negatively correlated (r=−0.51; Fig. 5).
3.5. Arsenic transfer across trophic levels Arsenic concentration in aquatic biota was higher in specimens sampled in the impacted than in the reference stream, and this difference was consistent across trophic levels (Table 2). Among primary producers, bryophytes had the highest accumulation (1,760 ± 485 μg g–1; mean ± SE), followed by the biofilm (449 ± 119 μg g–1) and macrophytes (29 ± 11 μg g–1; Fig. 4). Among aquatic invertebrates, arsenic concentration was highest in shredders (313 ± 139 μg g–1), followed by collectors (140 ± 58 μg g–1), predators (70 ± 20 μg g–1),
4. Discussion We investigated arsenic bioaccumulation by stream biota, and its 135
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4.2. Arsenic levels in the body reflect environmental levels Within the ranges of arsenic concentrations tested, stream biota accumulated arsenic proportionally to the arsenic concentration in the water, as reported by other investigators (Wang and Mulligan, 2006) and for other metals (e.g. Beghelli et al., 2016). Bryophytes and biofilm accumulated the most arsenic in our study, which is consistent with previous reports (Barral-Fraga et al., 2016) and supports the suggestions that (1) these organisms can be used as bioindicators of arsenic contamination (e.g. Favas et al., 2012; Varun et al., 2012) and (2) the ingestion of contaminated biofilm and other plant material is a possible route of arsenic flow into food webs. 4.3. Leaf-litter consumers acquire arsenic mainly by food ingestion Consumers can accumulate metals and trace elements by two routes, adsorption on the body surface or ingestion of food. Adsorption from the water is considered a main source of accumulation for invertebrates in metal-impacted streams (Rahman et al., 2012; Rainbow et al., 2007). However, Zubrod et al. (2015) found that most of the copper accumulated by Gammarus was obtained from food, and Schaller et al. (2009) found that Gammarus pulex accumulates uranium through both the body surface and food. We found that arsenic incorporation by the shredder S. vittatum occurred almost exclusively by food ingestion. Given the high capacity of biofilms to adsorb arsenic and the large surface: volume ratio of fine particulate organic matter (FPOM; a food source for many stream invertebrates), which facilitates arsenic adsorption, it is likely that other shredders, scrapers and FPOM consumers could also acquire arsenic from food. Larvae fed arsenic-contaminated leaf discs accumulated about 20-fold more arsenic than the larvae fed uncontaminated leaf discs (up to ~35 µg g−1), which is above the range of arsenic accumulation levels reported for trichopterans (0.002–0.18 µg g–1) by Wang and Mulligan (2006). If these laboratory findings can be extrapolated to environmental conditions, our study showed that arsenic contamination with levels as low as 40 µg L–1 can be magnified by detritus consumers. An unexpected finding was the faster growth of specimens fed arsenic-contaminated leaves, compared to those fed uncontaminated leaves. A reduction in growth with increased metal concentration has been reported for many consumers, including snails exposed to copper (Das and Khangarot, 2011; Mountain, 2015; Ng et al., 2011) and Heptagenia and Chironomus exposed to copper and cadmium (Milani et al., 2003). Increased growth of Cinygmula (Ephemeroptera) nymphs in a stream with moderate zinc pollution compared to reference streams was also reported by Carlisle and Clements (2005), although the authors explained the difference as differences in population life histories. The rapid growth observed for specimens fed arsenic-contaminated leaves could be an indirect hormesis effect on fungi, stimulating growth and resulting in increased maceration and better quality of litter, allowing shredders to grow rapidly (e.g. Calabrese and Baldwin, 2001). This possibility needs further investigation.
Fig. 4. Arsenic concentration (mean ± SE) in animals grouped in trophic levels, emergent adult insects and spiders in the reference and impacted streams. Biof: Biofilm; Bry: Bryophytes; Macr: Macrophytes; Scr: Scrapers; Col: Collectors; Fil: Filter-feeders; Shr: Shredders; Pred: Aquatic predators; Adult: Adult insects.
Fig. 5. Relationship between the arsenic concentrations and δ15N in different invertebrate trophic groups.
mobility in aquatic food chains and transfer to the terrestrial environment. We found that (1) low levels of arsenic caused changes in the structure and composition of the aquatic invertebrate assemblages studied; (2) arsenic accumulation by aquatic biota was related to arsenic in the environment; (3) leaf-litter consumers acquired arsenic mainly by food ingestion; and (4) there was no biomagnification of arsenic across trophic levels, and little arsenic was transported to the terrestrial environments by insect emergence. 4.1. Arsenic effects on aquatic invertebrate assemblages As hypothesized, at the site with high arsenic concentration, invertebrate density and richness were lower and the community composition differed from the reference site, an effect previously reported for arsenic (e.g. Chaffin et al., 2005) and for other metals (Iwasaki and Ormerod, 2012; Quinn et al., 2003). An important point of our study was that the effect was observed at a relatively low arsenic concentration (~40 µg L–1). This level was below the “no-effect concentration” (NOEC) reported in other aquatic invertebrates (~44 – 16,430 µg L–1) (Brix et al., 2003; Tišler and Zagorc-Končan, 2002) and is below the 150 µg L–1 US EPA (1999) levels for protection of aquatic life. Although arsenic toxicity depends on its species (Magalhães et al., 2015) and on the sensitivity of the taxa, our main conclusion regarding toxicity was that relatively low levels of arsenic were able to affect the aquatic biota.
4.4. Arsenic is not dispersed to terrestrial environments by insect emergence Our results suggest that arsenic accumulated by primary consumers is not magnified through the food webs, nor is it biologically dispersed from water to land. These findings agree with other studies (Juncos et al., 2016; Rahman et al., 2012; Watanabe et al., 2008). The decrease in arsenic concentration along trophic levels may be the result of efficiency in eliminating the contaminant (Watanabe et al., 2008). For instance, Mogren et al. (2013) found that Chironomus riparius excretes arsenic to the exoskeleton between the larval and pupal stages, while Schaller et al. (2015) observed that Gammarus pulex accumulates arsenic in the cuticle. If arthropods deposit arsenic in the exoskeleton, subsequent ecdyses will eliminate the arsenic from food webs. If the 136
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553–568. http://dx.doi.org/10.1899/04-058.1. Cui, B., Zhang, Q., Zhang, K., Liu, X., Zhang, H., 2011. Analyzing trophic transfer of heavy metals for food webs in the newly-formed wetlands of the Yellow River Delta, China. Environ. Pollut. 159, 1297–1306. http://dx.doi.org/10.1016/j.envpol.2011.01.024. Culioli, J.L., Fouquoire, A., Calendini, S., Mori, C., Orsini, A., 2009. Trophic transfer of arsenic and antimony in a freshwater ecosystem: a field study. Aquat. Toxicol. 94, 286–293. http://dx.doi.org/10.1016/j.aquatox.2009.07.016. Das, S., Khangarot, B.S., 2011. Bioaccumulation of copper and toxic effects on feeding, growth, fecundity and development of pond snail Lymnaea luteola L. J. Hazard. Mater. 185, 295–305. http://dx.doi.org/10.1016/j.jhazmat.2010.09.033. Doi, H., Takagi, A., Kikuchi, E., 2007. Stream Macroinvertebrate Community Affected by Point-Source Metal Pollution. Int. Rev. Hydrobiol. 92, 258–266. http://dx.doi.org/ 10.1002/iroh.200610923. Favas, P.J., Pratas, J., Prasad, M.N., 2012. Accumulation of arsenic by aquatic plants in large-scale field conditions: opportunities for phytoremediation and bioindication. Sci. Total Environ. 433, 390–397. http://dx.doi.org/10.1016/ j.scitotenv.2012.06.091. Feio, M.J., Graça, M.A.S., 2000. Food consumption by the larvae of Sericostoma vittatum (Trichoptera), an endemic species from the Iberian Peninsula. Hydrobiologia 439, 7–11. http://dx.doi.org/10.1023/A:1004189316952. Gaete, H., Chávez, C., 2008. Evaluación de la toxicidad de mezclas binarias de cobre, zinc y arsénico sobre Daphnia obtusa (Kurz, 1874) (Cladocera, Crustacea). Limnetica 27, 1–10. Gall, J.E., Boyd, R.S., Rajakaruna, N., 2015. Transfer of heavy metals through terrestrial food webs: a review. Environ. Monit. Assess. 187, 201–222. http://dx.doi.org/ 10.1007/s10661-015-4436-3. Ghaeni, M., Pour, N.A., Hosseini, M., 2015. Bioaccumulation of polychlorinated biphenyl (PCB), polycyclic aromatic hydrocarbon (PAH), mercury, methyl mercury, and arsenic in blue crab Portunus segnis from Persian Gulf. Environ. Monit. Assess. 187, 253–262. http://dx.doi.org/10.1007/s10661-015-4459-9. Graça, M.A.S., Cressa, C., Gessner, M.O., Feio, M.J., Callies, K.A., Barrios, C., 2001. Food quality, feeding preferences, survival and growth of shredders from temperate and tropical streams. Freshwat. Biol. 46, 947–957. http://dx.doi.org/10.1046/j.13652427.2001.00729.x. Ikemoto, T., Tu, N.P., Okuda, N., Iwata, A., Omori, K., Tanabe, S., Tuyen, B.C., Takeuchi, I., 2008. Biomagnification of trace elements in the aquatic food web in the Mekong Delta, South Vietnam using stable carbon and nitrogen isotope analysis. Arch. Environ. Contam. Toxicol. 54, 504–515. http://dx.doi.org/10.1007/s00244-0079058-5. Iwasaki, Y., Ormerod, S.J., 2012. Estimating safe concentrations of trace metals from inter-continental field data on river macroinvertebrates. Environ. Pollut. 166, 182–186. http://dx.doi.org/10.1016/j.envpol.2012.03.028. Juncos, R., Arcagni, M., Rizzo, A., Campbell, L., Arribere, M., Guevara, S.R., 2016. Natural origin arsenic in aquatic organisms from a deep oligotrophic lake under the influence of volcanic eruptions. Chemosphere 144, 2277–2289. http://dx.doi.org/ 10.1016/j.chemosphere.2015.10.092. Kehrig, H.A., Seixas, T.G., Malm, O., Di Beneditto, A.P., Rezende, C.E., 2013. Mercury and selenium biomagnification in a Brazilian coastal food web using nitrogen stable isotope analysis: a case study in an area under the influence of the Paraiba do Sul River plume. Mar. Pollut. Bull. 75, 283–290. http://dx.doi.org/10.1016/ j.marpolbul.2013.06.046. Liu, C.P., Luo, C.L., Gao, Y., Li, F.B., Lin, L.W., Wu, C.A., Li, X.D., 2010. Arsenic contamination and potential health risk implications at an abandoned tungsten mine, southern China. Environ. Pollut. 158, 820–826. http://dx.doi.org/10.1016/ j.envpol.2009.09.029. Magalhães, D.P., Marques, M.R.C., Baptista, D.,F., Buss, D.F., 2015. Metal bioavailability and toxicity in freshwaters. Environ. Chem. Lett. 13, 69–87. http://dx.doi.org/ 10.1007/s10311-015-0491-9. Marsden, I.D., Rainbow, P.S., 2004. Does the accumulation of trace metals in crustaceans affect their ecology — the amphipod example? J. Exp. Mar. Biol. Ecol. 300, 373–408. http://dx.doi.org/10.1016/j.jembe.2003.12.009. Mayer-Pinto, M., Underwood, A.J., Tolhurst, T., Coleman, R.A., 2010. Effects of metals on aquatic assemblages: what do we really know? J. Exp. Mar. Biol. Ecol. 391, 1–9. http://dx.doi.org/10.1016/j.jembe.2010.06.013. Milani, D., Reynoldson, T.B., Borgmann, U., Kolasa, J., 2003. The relative sensitivity of four benthic invertebrates to metals in spiked-sediment exposures and application to contaminated field sediment. Environ. Toxicol. Chem. 22, 845–854. http:// dx.doi.org/10.1002/etc.5620220424. Mogren, C.L., von Kiparski, G.R., Parker, D.R., Trumble, J.T., 2012. Survival, reproduction, and arsenic body burdens in Chironomus riparius exposed to arsenate and phosphate. Sci. Total Environ. 425, 60–65. http://dx.doi.org/10.1016/ j.scitotenv.2012.03.009 (DOI: 10.1016/j.scitotenv.2012.03.009). Mogren, C.L., Walton, W.E., Parker, D.R., Trumble, J.T., 2013. Trophic transfer of arsenic from an aquatic insect to terrestrial insect predators. PLoS One 8, e67817. http:// dx.doi.org/10.1371/journal.pone.0067817. Mountain, R.I.E.W.G., 2015. Elevation-dependent warming in mountain regions of the world. Nat. Clim. Change 5, 424–430. http://dx.doi.org/10.1038/NCLIMATE2563. Ng, T.Y.T., Pais, N.M., Wood, C.M., 2011. Mechanisms of waterborne Cu toxicity to the pond snail Lymnaea stagnalis: physiology and Cu bioavailability. Ecotoxicol. Environ. Saf. 74, 1471–1479. http://dx.doi.org/10.1016/j.ecoenv.2011.03.010. Painter, K.J., Janz, D.M., Jardine, T.D., 2016. Bioaccumulation of mercury in invertebrate food webs of Canadian Rocky Mountain streams. Freshw. Sci. 35, 1248–1262. http:// dx.doi.org/10.1086/688897. Qu, X., Wu, N., Tang, T., Cai, Q., Park, Y.S., 2010. Effects of heavy metals on benthic macroinvertebrate communities in high mountain streams. Ann. Limnol. - Int. J. Limnol. 46, 291–302. http://dx.doi.org/10.1051/limn/2010027.
process is efficient, then arsenic decreases through food webs, and adults may transport little arsenic when leaving the aquatic environment. At the arsenic levels observed in the stream (40 μg L–1), the mechanism for arsenic elimination was efficient, since we observed no differences in arsenic concentration in spiders living over the arsenic-impacted and reference streams. If spiders are good models of terrestrial consumers, we can conclude that there is no biodispersal of arsenic from water to land. 5. Conclusions At very low concentrations (~40 μg L–1), arsenic affected invertebrate assemblages, with changes in composition and taxa richness. Arsenic in streams is adsorbed by producers and by dead organic matter proportionally to its concentration in the water, and in this way enters food webs. Once incorporated into the producers, it can pass to consumers, but it is not biomagnified. Apparently, invertebrates have efficient mechanisms for arsenic elimination. It is possible for arsenic to be locked in the exoskeleton and eliminated during the ecdysis. Therefore, aquatic organisms do not cause the biodispersal of arsenic to terrestrial systems. Acknowledgements The authors thank Igor Moraes, Joana Sotaia Campos and Cristina Docal for their help in the field and laboratory work. Seena Sahadevan critically reviewed the manuscript. Rodrigo Fornel for the graphic abstract edition. LUH received support from the Programa Ciências Sem Fronteiras/CNPq/Brazil (Proc. #250080/2013-3). This research was supported by the Marine and Environmental Sciences Centre (MARE) through the program UID/MAR/04292/2013. Appendix A. Supporting information Supplementary data associated with this article can be found in the online version at doi:10.1016/j.ecoenv.2017.01.035. References Alonso, D.L., Latorre, S., Castillo, E., Brandão, P.F.B., 2014. Environmental occurrence of arsenic in Colombia: a review. Environ. Pollut. 186, 272–281. http://dx.doi.org/ 10.1016/j.envpol.2013.12.009. Ancion, P.Y., Lear, G., Lewis, G.D., 2010. Three common metal contaminants of urban runoff (Zn, Cu & Pb) accumulate in freshwater biofilm and modify embedded bacterial communities. Environ. Pollut. 158, 2738–2745. http://dx.doi.org/10.1016/ j.envpol.2010.04.013. Anderson, M.J., 2001. A new method for non-parametric multivariate analysis of variance. Austral Ecol. 26, 32–46. http://dx.doi.org/10.1111/j.14429993.2001.01070.pp.x. Awrahman, Z.A., Rainbow, P.S., Smith, B.D., Khan, F.R., Bury, N.R., Fialkowski, W., 2015. Bioaccumulation of arsenic and silver by the caddisfly larvae Hydropsyche siltalai and H. pellucidula: a biodynamic modeling approach. Aquat. Toxicol. 161, 196–207. http://dx.doi.org/10.1016/j.aquatox.2015.01.004. Barral-Fraga, L., Morin, S., Rovira, M.D.M., Urrea, G., Magellan, K., Guasch, H., 2016. Short-term arsenic exposure reduces diatom cell size in biofilm communities. Environ. Sci. Pollut. Res. 23, 4257–4270. http://dx.doi.org/10.1007/s11356-0154894-8. Beghelli, F.G.S., Pompêo, M.L.M., Rosa, A.H., Moschini-Carlos, V., 2016. Effects of copper in sediments on benthic macroinvertebrate communities in tropical reservoirs. Limnetica 35, 103–116. Brix, K.V., Cardwell, R.D., Adams, W.J., 2003. Chronic toxicity of arsenic to the Great salt Lake brine shrimp, Artemia franciscana. Ecotoxicol. Environ. Saf. 54, 169–175. Calabrese, E.J., Baldwin, L.A., 2001. Hormesis: a generalizable and unifying hypothesis. Crit. Rev. Toxicol. 31, 353–424. http://dx.doi.org/10.1080/20014091111730. Canivet, V., Chamnon, P., Gilbert, J., 2001. Toxicity and bioaccumulation of Arsenic and Chromium in epigean and hypogean freshwater macroinvertebrates. Arch. Environ. Contam Toxicol. 40, 345–354. http://dx.doi.org/10.1007/s002440010182. Carlisle, D.M., Clements, W.H., 2005. Leaf litter breakdown, microbial respiration and shredder production in metal-polluted streams. Freshw. Biol. 50, 380–390. http:// dx.doi.org/10.1111/j.1365-2427.2004.01323.x. Chaffin, J.L., Valett, H.M., Webster, J.R., 2005. Influence of elevated As on leaf breakdown in an Appalachian headwater stream. J. N. Am. Benthol. Soc. 24,
137
Ecotoxicology and Environmental Safety 139 (2017) 132–138
L.U. Hepp et al.
des eaux douces (Systématique élémentaire et aperçu écologique). Université Lyon, Lyon. Tišler, T., Zagorc-Končan, J., 2002. Acute and chronic toxicity of arsenic to some aquatic organisms. Bull. Environ. Contam. Toxicol. 69, 421–429. http://dx.doi.org/10.1007/ s00128-002-0079-5. Varun, M., D’Souza, R., Pratas, J., Paul, M.S., 2012. Metal contamination of soils and plants associated with the glass industry in North Central India: prospects of phytoremediation. Environ. Sci. Pollut. Res. 19, 269–281. http://dx.doi.org/ 10.1007/s11356-011-0530-4. Vithanage, M., Dabrowska, B.B., Mukherjee, A.B., Sandhi, A., Bhattacharya, P., 2011. Arsenic uptake by plants and possible phytoremediation applications: a brief overview. Environ. Chem. Lett. 10, 217–224. http://dx.doi.org/10.1007/s10311011-0349-8. Wang, S., Mulligan, C.N., 2006. Occurrence of arsenic contamination in Canada: sources, behavior and distribution. Sci. Total Environ. 366, 701–721. http://dx.doi.org/ 10.1016/j.scitotenv.2005.09.005. Wang, X., Peng, B., Tan, C., Ma, L., Rathinasabapathi, B., 2015. Recent advances in arsenic bioavailability, transport, and speciation in rice. Environ. Sci. Pollut. Res. 22, 5742–5750. http://dx.doi.org/10.1007/s11356-014-4065-3. Watanabe, K., Monaghan, M.T., Takemond, Y., Omura, T., 2008. Biodilution of heavy metals in a stream macroinvertebrate food web: evidence from stable isotope analysis. Sci. Total Environ. 394, 57–67. http://dx.doi.org/10.1016/ j.scitotenv.2008.01.006. Zubrod, J.P., Englert, D., Rosenfeldt, R.R., Wolfram, J., Lüderwald, S., Wallace, D., Schnetzer, N., Schulz, R., Bundschuh, M., 2015. The relative importance of dietrelated and waterborne effects of copper for a leaf-shredding invertebrate. Environ. Pollut. 205, 16–22.
Quinn, M.R., Feng, X., Folt, C.L., Chamberlain, C.P., 2003. Analyzing trophic transfer of metals in stream food webs using nitrogen isotopes. Sci. Total Environ. 317, 73–89. http://dx.doi.org/10.1016/s0048-9697(02)00615-0. Rahman, M.A., Hasegawa, H., Lim, R.P., 2012. Bioaccumulation, biotransformation and trophic transfer of arsenic in the aquatic food chain. Environ. Res. 116, 118–135. http://dx.doi.org/10.1016/j.envres.2012.03.014. Rainbow, P.S., Amiard, J.C., Amiard-Triquet, C., Cheung, M., Zhang, L., Zhong, H., Wang, W., 2007. Trophic transfer of trace metals: subcellular compartmentalization in bivalve prey, assimilation by a gastropod predator and in vitro digestion simulations. Mar. Ecol. Prog. Ser. 348, 125–138. http://dx.doi.org/10.1016/ j.envint.2006.05.007. Rhea, D.T., Harper, D.D., Farag, A.M., Brumbaugh, W.G., 2006. Biomonitoring in the Boulder River watershed, Montana, USA: metal concentrations in biofilm and macroinvertebrates, and relations with macroinvertebrate assemblage. Environ. Monit. Assess. 115, 381–393. http://dx.doi.org/10.1007/s10661-006-7086-7. Schaller, J., Brackhage, C., Dudel, E.G., 2009. Limited transfer of uranium to higher trophic levels by Gammarus pulex L. in contaminated environments. J. Environ. Monit. 11, 1629–1633. http://dx.doi.org/10.1039/b906420f. Schaller, J., Kochb, I., Caumetteb, G., Nearingb, M., Reimerb, K.J., Planer-Friedricha, B., 2015. Strategies of Gammarus pulex L. to cope with arsenic – results from speciation analyses by IC-ICP-MS and XAS micro-mapping. Sci. Total Environ. 530–531, 430–433. http://dx.doi.org/10.1016/j.scitotenv.2015.06.015. Subhani, M., Mustafa, I., Alamdar, A., Katsoyiannis, I.A., Ali, N., Huang, Q., Peng, S., Shen, H., Eqani, S.A., 2015. Arsenic levels from different land-use settings in Pakistan: bio-accumulation and estimation of potential human health risk via dust exposure. Ecotoxicol. Environ. Saf. 115, 187–194. http://dx.doi.org/10.1016/ j.ecoenv.2015.02.019. Tachet, H., Bournaud, M., Richoux, P., 1987. Introduction à l′étude des macroinvertébrés
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