Does road salting confound the recovery of the microcrustacean community in an acidified lake?
Introduction
The structure and function of freshwater ecosystems are affected by a multitude of anthropogenic stressors (e.g. Palmer and Yan, 2013). One such stressor is acidification of rivers and lakes, which has been a major environmental problem in North America and Europe for more than 50 years (Schindler et al., 1989, Skjelkvåle et al., 2005). Atmospheric deposition of acid components has significantly impacted water quality, resulting in marked decreases in pH and alkalinity and increases in concentrations of sulfate (SO4) and a subsequent mobilization of toxic aluminum (Al). Impoverished water quality has been followed by reduced diversity of aquatic organisms, e.g. a general shift in invertebrate assemblages (Larsen et al., 1996, Sandin et al., 2004, Schartau et al., 2008) and the loss of a vast number of fish populations (Gjedrem and Rosseland, 2012). As a result of international agreements and actions, the emission of sulfur (S) has been reduced by more than two-thirds since 1980 in Europe (EMEP, 2011) followed by improvements in surface water quality (Monteith et al., 2014, Skjelkvåle et al., 2005, Stoddard et al., 1999). Although recent studies have attributed recovery of lake biology to decreased deposition of acidifying compounds (Hesthagen et al., 2011, Raddum et al., 2004), evidence of widespread biotic recovery is generally lacking (Gray and Arnott, 2009, Gunn and Sandoy, 2003, Skjelkvale et al., 2001). Studies performed in Europe and North America have demonstrated a delay in zooplankton recovery of 3–10 years, even after water quality has reached acceptable levels (see Gray and Arnott, 2009). Thus, acidification is still considered a serious threat to the biodiversity and ecosystem functioning of inland surface waters in northern Europe (e.g. Brodin, 1995, EMEP The European Monitoring Evaluation Programme, 2005, EMEP The European Monitoring Evaluation Programme, 2011).
Biological recovery from acidification is a complex process. A number of biotic and abiotic factors may delay the restitution of the biological communities. Species interactions may prevent or delay the biological response to improved water quality (see Lundberg et al., 2000). For example, acid sensitive grazers may be excluded from their former niche by acid-tolerant generalists that have occupied this role during the period of acidification. Suppression by invertebrate predators may cause failure of herbivorous zooplankton to respond to the improved chemistry of acidified fishless lakes (Arnott et al., 2006). In other lakes, the return of and the predation by fish may impact on biological recovery (e.g. Valois et al., 2010). Dispersal limitation may also constrain or delay the recovery of biological communities (Gray and Arnott, 2011). An example of an abiotic factor that can delay biological recovery is the mobilization and leaching of dissolved organic carbon (DOC, Korosi and Smol, 2012), observed in many lakes of the Northern Hemisphere during the last decades and which may reduce the water clarity. Another example is the declining level of base cations as a consequence of long term acidification, which might affect certain crustacean zooplankton species (Alstad et al., 1999, Hessen and Rukke, 2000).
Increasing salt concentrations may also affect biological recovery from acidification. Higher salt concentrations of surface waters can be caused by deposits from the sea or from runoff of de-icing chemicals from roads. Here we will focus on the latter. The by far most applied de-icing chemical is sodium chloride (hereafter termed road salt). In the Northern Hemisphere sodium (Na) and chloride (Cl) concentrations of surface waters have increased over the last decades due to increasing use of road salt (e.g. Evans and Frick, 2001, Molot and Dillon, 2008). Also, in Norway there has been a significant increase in the amount of applied road salt, and since year 2000 the annual amount has more than tripled (Kronvall, 2013) leading to increasing Cl concentrations in lakes (Bækken and Haugen, 2012).
Increasing salt concentration in lakes affects the ecological structure and functioning of the aquatic ecosystem both directly and indirectly. At high enough concentrations, Cl can increase the acidity of water, causing some of the same negative effects as acid rain (Löfgren, 2001). Higher salt concentrations may also favor more salt tolerant taxa (Blinn et al., 2004, Evans and Frick, 2001, Sarma et al., 2006). However, one of the most evident impacts of road salt on the aquatic environment is increased density stratification followed by altered circulation patterns and oxygen depletion in lakes (Kjensmo, 1997, Koretsky et al., 2012, Novotny et al., 2008, Novotny and Stefan, 2012). A recent survey conducted in Norway revealed that almost 30% of the investigated lakes had developed salt gradients and altered circulation patterns (Bækken and Haugen, 2012). Oxygen depletion of the deep water may in turn lead to mobilization of toxic metals and phosphorus from the sediments (Bækken and Haugen, 2012, Lewis, 1999). Increased salt concentrations may therefore indirectly lead to algal blooms because of increasing nutrient availability. So far most of the literature dealing with the effects of increasing salt concentrations on aquatic freshwater ecosystems has focused on relatively high concentration ranges (e.g. Van Meter et al., 2011). However, Evans and Frick (2001) noted that very small increases in Cl of 2–10 mg L− 1 can affect phytoplankton community structure in low salinity lakes. In Norwegian calcium poor lakes effects on the phytoplankton community may appear at chloride concentrations of 23–30 mg L− 1 (Haugen et al., 2011). Growth reductions of individual sensitive phytoplankton species occur at even lower concentrations (10–15 mg L− 1, Haugen et al., 2011). A recent study of multiple anthropogenic stressors on Canadian freshwater zooplankton demonstrated effects of rising Cl concentrations on important zooplankton metrics at concentrations below 5 mg L− 1 (Palmer and Yan, 2013). The effect of Cl at low concentrations may be exacerbated in a multiple stressor environment, e.g. recovering from acidification.
Numerous lakes across the temperate and subarctic zones in North America and Northern Europe undergoing chemical recovery from acidification are also experiencing a simultaneous increase in Cl concentrations due to deicing of roads during winter (e.g. Palmer and Yan, 2013). No studies have investigated if the biological recovery from acidification is affected by the increasing Cl concentrations in these lakes. The Lake Øvre Jerpetjern in Southeastern Norway (Fig. 1) undergoing chemical recovery from acidification is also influenced by road salts. Using monitoring data on water chemistry and microcrustaceans from this lake, we test whether the change in water quality (decreasing acidification and increasing Cl concentrations) over a 17 year period has affected the structure of the microcrustacean community. A comparable lake recovering from acidification but without influence from road salts was used as a reference. We hypothesize that increasing Cl concentration in the lake modifies the recovery of the microcrustacean community from acidification.
Section snippets
Study site
Lake Øvre Jerpetjern is located in Southeastern Norway in the county of Telemark about 48 km from the sea (N 59.60692°, E 9.42168°). The lake is situated at an altitude of 457 m.a.s.l. It has a surface area of 0.12 km2, and the maximum depth is 17 m (Fig. 1). The catchment area is 1.85 km2, mainly covered by forest consisting of spruce (Picea abies) and pine (Pinus sylvestris). The bedrock of the catchment consists of granite. One of the main roads between Southeastern and Western Norway, the E134,
Water chemistry
Chemical monitoring of Lake Øvre Jerpetjern from 1986 showed that the increase in the concentration of nmSO4 was followed by a decrease since 1991 with a subsequent increasing trend in pH and decreasing trend in LAl. Also the concentration of TOC increased. Moreover, the concentration of Ca has decreased from 1995 (Fig. 2). The changes in the acidification related parameters indicate that the lake has undergone a period of chemical recovery. Concurrent with the improvement of the acidification
Discussion
Lake Øvre Jerpetjern has been going through a period of acidification followed by chemical recovery from the early 1990s onwards. Wright et al. (unpublished results) calculated the original water chemistry (pH and acid neutralizing capacity (ANC)) in the lake before acidification started (1800) using the F-factor method (Henriksen and Posch, 2001). Their results demonstrate that with regard to these parameters, the lake shows only a minor departure from pre-industrial conditions from 2004
Conclusions
In Lake Øvre Jerpetjern, experiencing an increase in chloride, the microcrustacean community is recovering from acidification. Our analysis shows that the reduced acidification pressure was the main driver of the changes of the community. However, it is evident that the recovery of the community was lagging behind the chemical recovery, and was slower than in an acidified “reference lake”, not subject to road salt. Recurrent episodes of low pH and high LAl and a decreasing Ca concentration are
Conflict of interest
The authors of this manuscript declare to have no conflict of interests.
Acknowledgments
The study was supported by the Norwegian Institute for Nature Research and the Norwegian Public Roads Administration. “Monitoring of long-range transboundary air pollution” was financed respectively by the Norwegian Climate and Pollution Agency (Klif) and the Norwegian Directorate for Nature Management (DN) (now Norwegian Environment Agency). Erik Framstad and three anonymous reviewers provided valuable suggestions and comments on the manuscript.
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