Variation in Zn, C, and N isotope ratios in three stream insects

With the tremendous advances in mass spectrometry (MS) techniques over the past few decades, the use of stable isotope ratios to determine the cycling and ultimate fate of elements in the aquatic environment has grown exponentially. Mass-dependent fractionation of lighter isotopes, such as C and N, has been well-studied. Variations in stable isotope ratios of ¹³C/¹²C (δ ¹³C) and ¹⁵N/¹⁴N (δ ¹⁵N) have been used extensively to examine food sources, trophic structure, and energy transfer and efficiency in lakes (Vander Zanden et al. 1999; Vander Zanden and Rasmussen 1999, 2001; Post 2002), reservoirs (Saito et al. 2001), and marine systems (Das et al. 2003). In addition, interest in the sources and trophic transfer of contaminants in aquatic systems has led to studies in which δ ¹⁵N and (or) δ ¹³C values were coupled with measurements of total metal concentrations in both freshwater and marine biota (e.g., Power et al. 2002; Das et al. 2003; Quinn et al. 2003; Larsson et al. 2007; Jardine et al. 2009, 2012; Senn et al. 2010; Jones et al. 2014).

Previous limitations in MS technology precluded rapid, precise, sensitive, and accurate measurements of stable isotope ratios of the transition elements and other heavier elements. However, developments in multicollector (MC)-inductively coupled plasma (ICP)-MS technology have enabled investigations of stable isotope ratios for transition metals, such as Fe, Cu, and Zn, and also heavier elements, such as Hg, similar to those for lighter elements (see review by Weiss et al. 2008). Thus, the potential exists to use transition element stable isotope ratios to trace kinetic pathways of these elements in biological systems, similar to the approach used for the lighter elements, C and N. To date, most work has been directed towards isotopic fractionation of Hg (a known toxin) in various environmental compartments (e.g., sediments, zooplankton, and fish), which, in combination with C and N stable isotope values, has been used to trace sources of Hg (Gantner et al. 2009; Perrot et al. 2010, 2012).

In humans, zinc is an essential nutrient that plays an important structural role in a number of proteins including those involved in metabolism, protein expression, and replication; however, acute and chronic zinc toxicities can occur at high Zn concentrations. Zinc also has several stable isotopes making it an ideal candidate for isotopic fractionation studies that can be extrapolated potentially to other transition metals. Zinc fractionation has been determined in a variety of materials (see review by Cloquet et al. 2008) including, for example, seawater (Conway et al. 2013), soil profiles (Juillot et al. 2011), and plants (Weiss et al. 2005; Cavagnaro and Jackson 2007; Viers et al. 2007). Results from terrestrial plant studies suggest that the lighter isotope is taken up preferentially or is more available (Deng et al. 2014), although enrichment of heavy Zn isotopes (δ ⁶⁶Zn = 0.40‰–0.72‰) was observed from soil to root, followed by a depletion (δ ⁶⁶Zn = −0.10‰ to −0.50‰) in heavy Zn from root to shoot (Tang et al. 2012).

Despite evidence that Zn fractionation occurs in geological material, soils, plants, and water (Cloquet et al. 2008), there are few data on whether this is also the case in animals. In sheep, Balter et al. (2010) found that bone, muscle, serum, and urine were enriched in the heavy Zn isotopes relative to the diet, whereas the feces, red blood cells, kidney, and liver were enriched in light isotopes. Jaouen et al. (2013) reported that for animals and plants collected from Western Cape and Kruger National Park, South Africa, the bones of herbivores were significantly enriched in the heavier Zn isotopes relative to the plants, whereas the carnivores were slightly depleted in the heavier isotopes compared to the herbivores. Average zinc concentrations were similar in the animals (117 ± 95 μg/g) and plants (74 ± 38 μg/g), and no correlation was observed between total Zn concentration and isotopic composition. To our knowledge, there are no data on Zn fractionation in aquatic insects.

Differences in δ ¹⁵N and δ ¹³C values between trophic levels and from different food sources, respectively, have been used repeatedly to assess food sources, trophic structure, and energy transfer and efficiency in freshwaters (Vander Zanden et al. 1999; Vander Zanden and Rasmussen 1999, 2001; Saito et al. 2001; Post 2002). Data collected from stream invertebrates indicated the potential for a significant difference in whole-body metal concentrations according to feeding strategy (Smock 1983). Thus, it is possible that Zn also may fractionate according to food sources or feeding strategy. Therefore, the purpose of this study was to measure Zn isotope ratios, δ ¹⁵N and δ ¹³C in three common stream insects, with different feeding strategies, collected from a stream in Ontario, Canada, to test whether Zn isotope fractionation occurs and if so, whether or not it is related to feeding strategy.

A total of 40 stonefly nymphs, 41 caddisfly larvae, and 42 water striders were collected on the 4 sampling days. Total Zn, δ ⁶⁶Zn, δ ¹⁵N, and δ ¹³C values were determined on individual insects, although each parameter was not always measured on each insect. As there was no difference in the average total Zn, δ ⁶⁶Zn, δ ¹⁵N, and δ ¹³C values in insects where all parameters had been measured compared to those where they had not, all data were pooled.

Average total Zn concentrations (Fig. 1a) were highest in the water striders (193 ± 88.5 μg Zn/g dw, n = 34) and were significantly higher (P < 0.05) than those measured in the stoneflies (136 ± 34 μg Zn/g dw, n = 38) and caddisflies (125 ± 26 μg Zn/g dw, n = 40), which were not significantly different from each other. Total Zn concentrations remained constant in all three insects, both as a function of time and also dw. Average dw of the stoneflies (0.045 ± 0.026 g) and the caddisflies (0.040 ± 0.010 g) were significantly higher than the average dw of the water striders (0.010 ± 0.005 g).

Zn fractionation in the stoneflies and caddisflies did not vary significantly over the 4-month sampling period; however, δ ⁶⁶Zn in water striders was significantly lower (P < 0.05) in July (−2.288‰ ± 0.662‰) than in May and August (Fig. 1b; note no δ ⁶⁶Zn values were determined on water striders collected in June). When the data were combined, ANOVA revealed that δ ⁶⁶Zn values for water striders (−1.196‰ ± 1.030‰) were significantly lighter (more negative) than those measured in the stoneflies (−0.453‰ ± 0.616‰) and caddisflies (−0.512‰ ± 0.542‰), which were similar to each other. There was no correlation between δ ⁶⁶Zn and total Zn concentrations for any of the insects (data not shown).

The results for C and N fractionation indicate a significant difference in δ ¹³C values between caddisflies and water striders (−30.682‰ ± 1.012‰ and −28.611‰ ± 0.981‰, respectively) and also between stoneflies and water striders (−30.754‰ ± 1.327‰ and −28.611‰ ± 0.981‰, respectively) but not between caddisflies and stoneflies (Fig. 2a). Water striders had the lowest and most variable δ ¹⁵N signature (6.810‰ ± 2.726‰; Fig. 2b), whereas stoneflies had the highest and least variable δ ¹⁵N signature (9.604‰ ± 0.396‰); caddisflies were intermediate (7.581‰ ± 0.513‰). There was a significant difference (P < 0.05) in δ ¹⁵N values among all three species, that is, between caddisflies and stoneflies, between caddisflies and water striders, and between water striders and stoneflies.

There was no correlation between δ ⁶⁶Zn and δ ¹³C or δ ⁶⁶Zn and δ ¹⁵N in any of the insects (data not shown).

The average concentration of total Zn in the water striders is significantly higher than in the stoneflies and caddisflies. Even if one very high value for total Zn in one of the water striders (i.e., 544 μg/g dw; Fig. 1a) is removed from the average, total Zn concentrations in the water striders (170 ± 44 μg/g dw) are still significantly higher than in the caddisflies (128 ± 23 μg/g dw) or stoneflies (141 ± 34 μg/g dw). This most likely represents bioaccumulation; water striders are predators and are ingesting and accumulating Zn from their (terrestrial) prey similar to what was observed by Quinn et al. (2003), who found that Zn concentrations increased with trophic level in a stream food web in Montana. In addition, lower Zn levels in the heavier caddisflies and stoneflies could reflect simple bio-dilution of the metal in these insects.

The Zn isotope ratios in the water striders indicate that the heavier isotopes are significantly depleted (more negative delta values) relative to those of the caddisflies and stoneflies, whereas there is no significant difference between caddisflies and stoneflies (Fig. 1b). The large variability and more negative δ ⁶⁶Zn values for water striders in July could reflect an overlap of nymph and adult stages, given that they can coexist during the summer (Blanckenhorn and Fairbairn 1995). Evidence for lighter δ ⁶⁶Zn signatures in predators was reported by Jaouen et al. (2013) who observed that ⁶⁶Zn/⁶⁴Zn was slightly depleted (by ∼0.2‰) in carnivore bones relative to herbivore bones. Depletion of ⁶⁶Zn from the roots to shoots of higher plants (lettuce, tomato, rice, and Durham wheat) under varying exposure conditions (e.g., nutrients and Fe levels; Weiss et al. 2005; Jouvin et al. 2012) and moving from root to shoot to fruit and leaves in a variety of plants has also been reported (Viers et al. 2007; Cavagnaro and Jackson 2007). Assuming that terrestrial herbivores feed on terrestrial plants, then heavier isotope depletion would also be expected in their predators. This, therefore, suggests that diet may account for the observed differences in Zn fractionation among the stream insects.

If the observed fractionation of Zn is due to dietary sources, we can further examine why the heavy isotopes of Zn are depleted in the water striders compared to the other groups. The original selection of stream insects was based on different feeding styles and, therefore, potential differences in diet. Pycnopsyche spp. are shredders with a preference for microbially conditioned detrital material (Rong et al. 1995; Collins et al. 2016). A. abnormis is a primarily predatory stonefly with prey selection dependent on availability of prey items (Genito and Kerans 1999; Benke et al. 2001). In Jackson Creek primary prey items would be mayflies and caddisflies.

In contrast, water striders are entirely carnivorous and have a significant range in food choice, consisting primarily of terrestrial insects that land on the water surface (Smock 1983) and to a lesser extent, autochthonous aquatic larvae that float on the surface. The average δ ¹³C value of −28.61‰ ± 0.98‰ observed in the water striders is within the range of values reported for A. remigis in New Brunswick, Canada (∼−26‰ to −32.5‰; Jardine et al. 2012) and is significantly heavier (i.e., less negative) than the δ ¹³C values measured in our stoneflies and caddisflies (Fig. 2a). Jardine et al. (2012) found that A. remigis was only weakly linked to aquatic organic matter, indicating allochthonous (terrestrial) prey sources for these insects; in an earlier publication, Jardine et al. (2009) even advocated that adult water striders were potentially more appropriate as terrestrial, rather than aquatic sentinels of Hg sources. Alternately, the similarity in average δ ¹³C values between the stoneflies (−30.754‰ ± 1.327‰) and caddisflies (−30.682‰ ± 1.012‰) indicates that they have comparable aquatic sources of dietary carbon. Jardine et al. (2012) found significant correlations between δ ¹³C_insect and δ ¹³C_periphyton, indicating a strong connection to autochthonous C sources, despite the fact that Pycnopsyche are known to shred terrestrial leaf litter. More recently, the importance of autochthonous bacterial and fungal growth on leaf litter as a dietary carbon source to Pycnopsyche and other stream invertebrates such as Ephemeroptera was documented (Chung and Suberkropp 2009; Halvorson et al. 2016; Collins et al. 2016). It appears that a primary source of C incorporated by stream insects, such as Pycnopsyche, comes from autochthonous bacteria and fungi that populate the terrestrial leaf litter. For this reason, their δ ¹³C signature is more reflective of streamwater ¹³C than that of terrestrial vegetation. Given that caddisflies and mayflies are a primary food source for the stoneflies, it, therefore, makes sense that the δ ¹³C signatures of the stoneflies and caddisflies are not significantly different. Thus, our C isotope ratio analyses corroborate the food source differences/similarities for these three organisms.

The lower δ ¹⁵N values in the water striders compared to the stoneflies and caddisflies are perhaps contrary to what would be predicted, given that they are entirely carnivorous. The predatory stoneflies have a slightly higher signature than the caddisflies, which are shredders. Generally, a ∼2.5‰ to 3.4‰ enrichment in δ ¹⁵N is assumed from diet to consumer (DeNiro and Epstein 1981; Post 2002). However, differences in diet-tissue ¹⁵N fractionation in stream invertebrates have been reported suggesting that ¹⁵N isotope signatures do not always follow this pattern in aquatic insects (Jardine et al. 2012). The method by which water striders feed can affect their δ ¹⁵N values. McCutchan et al. (2003) proposed that estimates of the trophic shift for nitrogen differ between “fluid-feeding” consumers (such as water striders, which suck out nourishment from their prey) and other consumers that eat the entire prey, so the lower δ ¹⁵N values we measured in the water striders might reflect this feeding behavior. Also, species which undergo complete metamorphosis (caddisflies) might be expected to have a longer or more highly processed or more variable food chain length, which would also modify δ ¹⁵N values. Thus, the differences in δ ¹⁵N values among the stoneflies, caddisflies, and water striders may indicate slight dietary differences and (or) different feeding strategies resulting from the species or instar present during collection of the organisms.

Borrok et al. (2008) observed that in streams with low dissolved Zn concentrations, the heavier isotopes were more abundant, suggesting that the heavy isotopes are bound preferentially by dissolved organic carbon (DOC) and micro-organisms such as the bacteria and fungi populating detrital material in the water. The similarity in δ ¹³C and δ ⁶⁶Zn between caddisfly larvae and stonefly nymphs suggests that the path of Zn incorporation and fractionation is similar to that of carbon. It is generally accepted that stream invertebrates take up the majority of their Zn from food (organic material) and not water (reviewed by Goodyear and McNeill 1999), further supporting the expectation of a heavier signature in these organisms relative to water striders. Water striders on the other hand, having a diet consisting primarily of terrestrial insects that would be feeding on plant parts depleted in the heavy isotopes, reflect their food source. Although we do not have isotope ratio data for terrestrial insects, given the correspondence between the C and Zn isotope ratios, it is reasonable to conclude that the differences observed between species relate to differences in Zn ratios in their diets.

In conclusion, our findings suggest that sources of Zn to these aquatic insects are similar to that of C; it may be possible to trace sources of Zn, and possibly other contaminants, in food webs using Zn isotope ratios in stream invertebrates with different feeding styles. Chen et al. (2008) used Zn isotope signatures to trace contaminants in river water; however, this required large volumes of water and extensive concentration of samples. Invertebrates already contain adequate concentrations of Zn for analysis, and, therefore, less sample preparation is required resulting in less chance of contamination. The use of water striders as a sentinel of Hg sources has already been proposed (Jardine et al. 2005). Furthermore, improvements in ICP-MS techniques have resulted in recent growth in studies where C, N, and Hg stable isotope ratios have been combined to address source partitioning and biogeochemical transformations of Hg in environmental compartments (Gantner et al. 2009; Perrot et al. 2010, 2012; Tsui et al. 2012).

Our findings contribute to the overall understanding of the biogeochemical cycling of Zn, specifically in freshwater communities. Future work will require analyzing a wider range of stream invertebrates for both Zn and C isotope fractionation, as well as analysis of specific food and water sources. By expanding the scope of the study, it should be possible to partition terrestrial and aquatic sources of Zn to biota, and from there, to identify all environmental sources (not simply food) of Zn and other transition metals.

This work was supported by an NSERC Strategic Grant to RDE. The authors thank Jordan Balch for laboratory and field assistance and three anonymous reviewers for their critical input.