Reclaimed Water Systems: Biodiversity Friend or ... - ACS Publications

Chapter 17

Downloaded by UNIV OF CALIFORNIA SAN DIEGO on December 8, 2015 | http://pubs.acs.org Publication Date (Web): December 3, 2015 | doi: 10.1021/bk-2015-1206.ch017

Reclaimed Water Systems: Biodiversity Friend or Foe? Wei Zhang,*,1 Christopher Saint,1 Philip Weinstein,2,3,4 and David Slaney3 1Centre

for Water Management and Reuse, University of South Australia, Mawson Lakes, SA 5095, Australia 2School of Pharmacy and Medical Sciences, University of South Australia, Adelaide, SA 5001, Australia 3Barbara Hardy Institute, University of South Australia, Adelaide, SA 5001, Australia 4School of Biological Sciences, University of Adelaide, Adelaide, SA 5005, Australia *E-mail: [email protected].

Surface-flow type constructed wetlands are being commonly prescribed by local authorities for stormwater and wastewater treatment in many parts of Australia, and elsewhere. Despite little is known about the biodiversity in constructed wetlands up to date, these artificial wetlands can potentially develop and become complex ecosystems with a variety of fauna and flora including macroinvertebrate communities. Such communities could play an important role in the ecological functioning of healthy aquatic ecosystems because they process detritus and algae, and also provide food to other aquatic animals. In the meantime, nuisance macroinvertebrate species such as chironomid midges and mosquitoes associated with man-made systems could also raise significant health concerns as disease vectors. Water quality conditions such as low dissolved oxygen, temperature, nutrients, pH, heavy metal and organic contaminants are the key determinants of macroinvertebrate communities in man-made reclaimed water systems. Therefore, the potential impact of different reclaimed water quality parameters on aquatic macroinvertebrate assemblage has been reviewed systematically. This chapter underscores the © 2015 American Chemical Society In Water Challenges and Solutions on a Global Scale; Loganathan, et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

importance of promoting biodiversity through the maintenance of healthy wetlands (i.e. water quality monitoring and control aspect) and beneficial and sustainable ecosystem services which have consequences for human health (i.e. disease vector regulation).


Introduction Water quality and availability is a critical issue facing Australia (1). In recent years in urban areas of Australia reclaimed water (i.e. treated wastewater and storm water) has been stored and used to supplement potable water use. This is seen as an extremely valuable resource, especially in times of drought. In addition, developers and local authorities are increasingly considering the amenity aspects of the systems that are used to store and transfer this water supply (such as wetland, pond and canal areas) and they are seen as adding aesthetic value to such developments. Ironically, little is known about the biodiversity of such systems and this gives rise to the following research questions: 1. 2.

Can they supplement stressed natural systems by providing refugia for aquatic organisms? (The “friendly” situation) Could they potentially be habitat for nuisance speciessuch as disease vector mosquitoes and biting midges, increasing the public health risk to the co-located communities? (The “foe” situation)

Macroinvertebrate assemblages have been used frequently for evaluating the status of aquatic ecosystems, as a complement to physical–chemical indicators, due to their ubiquity, diversity, and their wide range of tolerances to natural and human‐induced environmental variation (2). Reclaimed water, such as treated waste and urban storm water, can deliver a range of contaminants (i.e. heavy metals and organic pollutants) to receiving water systems, which may influence the macroinvertebrates present. Monitoring water quality in man-made reclaimed water systems and relating this to the macroinvertebrate types seen should provide an indication of the tolerances of a range of species to water of varying quality. This in turn will inform the future design and operation of such systems where optimal biodiversity outcomes are required. In this chapter, we will first discuss the biodiversity contribution of macroinvertebrate assemblages such systems may provide in tandem with assessing the potential health risks from nuisance species. We will then link the presence of these macroinvertebrate assemblages to different water quality parameters in the constructed reclaimed water systems. At the moment, there are very few studies on macroinvertebrate assemblages in man-made reclaimed water systems and the effects of reclaimed water quality on macroinvertebrate biodiversity. Due to the paucity of information on reclaimed water systems, studies of a similar nature on natural aquatic systems (i.e. river, lakes and natural wetlands) will be chosen in many cases and reviewed to shed some light on this aspect. This could provide guidance for future research in this field. 356 In Water Challenges and Solutions on a Global Scale; Loganathan, et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.


Reclaimed Water Systems as Potential Refugia for Macroinvertebrate Assemblages - The “Friendly” Situation Macroinvertebrates are organisms that are large (i.e. macro) enough to be seen with the naked eye and lack a backbone (i.e. invertebrate). They inhabit all types of running waters, from fast flowing mountain streams to slow moving muddy rivers. Examples of aquatic macroinvertebrates include insects in their larval, nymph, or adult form, as well as crayfish, clams, snails, mites, sponges and worms. Macroinvertebrates play an important role in the ecological functioning of healthy aquatic ecosystems because they process detritus and algae, and also provide food to other aquatic animals. The most common types of aquatic macroinvertebrates are insects. For example, Chironomidae (i.e. non-biting midges) are probably the most widely distributed and species rich family and can constitute between 10% and 50% of the biomass of aquatic macroinvertebrates (3). Due to their ubiquitous nature and sensitivity to water quality changes, some insect species are well documented as being good biological indicators in aquatic ecosystems, such as Ehpemeroptera (i.e. mayflies), Plecoptera (i.e. stoneflies), and Trichoptera (i.e. caddisflies) (EPT) (2). For most aquatic insects, their juvenile stage usualyoccupies the major proportion of their life cycle and the adulthood only plays a brief reproductive role. For example, some dragonfly larvae take three years to mature. Reclaimed water wetlands within Australian suburban areas are increasingly valued for aesthetic, biodiversity, flood and water management, and recreational reasons, as well as for providing habitat for birds. For these man-made reclaimed water systems, the bed may be sandy or muddy with increased light penetration due to the slow moving or still nature of the water (i.e. lentic). Nutrients are available and produce conditions for algal growth. Catergorized as diferent functional feeding groups, collectors and scrapers (such as mayfly nymph, mussels, water fleas, some fly larvae and worms) are more likely to dominate the macroinvertebrate community. Whereas collectors will burrow into the sediment or filter their food directly from the water column, scrapers/grazers (such as snails, limpet and mayfly larvae) will be found on rocks, snags and woody debris or aquatic plants. It has been shown that the composition and diversity of macroinvertebrate community can be enhanced by increasing the amount of riparian vegetation and aquatic plants along streams and upland areas (4). Their presence can balance the light availability, reduce extreme temperature, protect banks from erosion, and act as nutrient filters.

Potential Human Health Risks Associated with Reclaimed Water Systems - The “Foe” Situation A diverse range of bacterial and viral waterborne pathogens are present in urban sewage and agricultural run-off. Wastewater is treated to a high standard before it is permitted to be used for re-use purposes and a range of guidelines relate quality to permitted usage, sometimes referred to as “fit for purpose” (Australian Guidelines for Water Recycling, 2007). The collection of storm 357 In Water Challenges and Solutions on a Global Scale; Loganathan, et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.


water is not without significant risk in this regard, as catchments often contain sewers that overflow or leak during intense weather events. The list of potentially pathogenic viruses commonly present in urban wastewater treatment plants include the DNA viruses, adenovirus and polyomavirus, and RNA viruses, such as enterovirus, hepatitis A and E viruses, norovirus, rotavirus and astrovirus (5). Common waterborne bacterial pathogens present in reclaimed water include faecal coliforms, E. coli, Enterococci, Salmonella, Campylobacter, Listeria monocytogenes, Vibirovulnificus, Helicobacter pylori, etc (5). Protozoan parasites (around 0.01 mm to 1.0 mm in length) is another group of waterborne pathogens and have been detected from various sources of reclaimed water, including Cryptosporidium, Giardia, Cyclospora, Toxoplasma, Microsporidia, Entameba, Acanthamoeba spp., Isospora, Blastocystishominis, Sarcocystis spp., Naegleria spp. and Balantidium coli, etc (5). Mosquitoes (Culicidae family) are a diverse taxonomic group that plays a number of important roles in healthy wetlands, where they can interact with a variety of invertebrates and vertebrates in complex communities within wetlands. They usually prefer water bodies that are shallow, lentic and warm. Mosquitoes are also capable of acting as vectors of human pathogens of various types, such as protozoa (e.g. malaria), nematodes (filaria) and viruses (‘arboviruses’: a contraction of ‘arthropod borne-virus’, e.g. Ross River virus, dengue, West Nile virus, and yellow fever), although only a small proportion of mosquito species are known vectors for human disease. Constructed wetlands are of concern because densities of larval populations in constructed wetlands can greatly exceed those in natural wetlands (6). It is difficult to generalise the range and number of species that will exploit constructed wetlands in Australia and elsewhere. Particular species that are of concern for reasons of greater pest or vector potential in Australia are summarised in Table 1, although there are similar situations in other countries where constructed wetlands are being used for wastewater treatment in regions with mosquito/wetland/pathogen associations (7, 8). Biting midges (Ceratopogonidae family) are another insect group that are of concern in relation to constructed wetlands. Of the 70 plus genera of biting midges, only four contain species that feed on vertebrates including humans, the most important genus being the Culicoides (9). Unlike mosquitoes, biting midges vector only a small number of diseases and parasites to humans and are not known to vector any of these to humans in Australia. They are primarily a pest species from their biting, which can induce severe local allergic reactions particularly in children. Biting midges occur in a wide range of aquatic and semiaquatic habitats, including wetlands, where their immature stages develop in areas of nutrient richness (9). Given this association with nutrient rich substrates, reclaimed water systems that deliver excess nutrients e.g. via urban stormwater runoff, can lead to increasesin biting midge populations. An assessment of the risks presented by mosquitoes in reclaimed water wetlands can be of great complexity, which is dependent on the mosquito species present, their access to pathogens, environmental conditions, and contact with humans. There are several deciding factors on whether a particular species is a local vector of a pathogen: (i) physiological factors that determine whether the mosquito can be infected with, maintain, and transmit the pathogen after an 358 In Water Challenges and Solutions on a Global Scale; Loganathan, et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.


incubation period, and (ii) behavioural and environmental factors that determine whether the species is sufficiently abundant and has sufficient contact with the reservoir host of the pathogen and later with humans, and will live long enough in sufficient numbers to transmit the pathogen. A particular species may be more important in one area than another because of variations in the above features between populations. Despite the various factors involved, mosquito-borne disease is primarily density dependent; the major factor in transmission risk is mosquito abundance, and this can be evaluated by sampling the mosquito populations (14). For example, inspections for mosquito larvae in the habitat can give an indication of the presence of mosquito populations and site-specific ‘threshold levels’ can be provided (e.g. Breteau index (15)).

Table 1. Pest or Potential Disease Vectoring Mosquito Species Likely to Be Found Associated with Various Constructed Non-Tidal Wetlands in Australia Genus

Main species of concern

References

Anopheles

Range of Anopheles species are potential malaria vectors e.g. Anopheles annulipes, An. farauti and An. bancroftii (malaria was declared erradicated in Australia in 1981).

(10, 11)

Culex

Culexannulirostris, vector for Murray Valley encephalitis, Kunjin, Ross River and Barmah Forest viruses

(12, 13)

Coquillettidia

Coquillettidialinealis, predominantly a bitting pest species and a potential vector for Ross River and Barmah Forest viruses; and Cq. xanthogaster, predominantly a bitting pest species

(11, 13)

Mansonia

Mansoniauniformis, predominantly a bitting pest species

(10, 11, 13)

Aedes

Aedesbancroftianus, Ae. sagax, Ae. theobaldi and Ae. vittiger,predominantly bitting pest species; and Ae.camptorhynchus and Ae. vigilax (saline influenced wetlands), vectors for Ross River and Barmah Forest viruses

(9, 11, 12)

359 In Water Challenges and Solutions on a Global Scale; Loganathan, et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Effects of Water Quality Parameters on Macroinvertebrate Assemblages in Reclaimed Water Systems


Temperature Reclaimed water from industrial discharges or urban storm water runoff usually has elevated water temperatures (16, 17), probably as a result of being heated by impervious surfaces (e.g. roads and car parks). The wider and more open channels of incised urban streams could probably contribute to an increase in the range of temperature variation between day and night. Warmer water is likely to stimulate physiological processes in streams and aggravate the problems of nuisance algal growth.Water with higher temperature also holds less dissolved oxygen (DO) than cooler water and forms an environment potentially less beneficial for macroinvertebrate communities.Some macroinvertebrates that are adapted to cool waters, such as stoneflies, are likely to suffer thermal stress in warm water. Some studies have demonstrated that elevated temperature (within 15°C to 25°C range) of the downstream of small farm dams show little effect on the overall macroinvertebrate communities (i.e. EPT richness) (18), while other studies suggest that their diversity tends to increase with increasing temperature (19). Similar effects are also likely to apply to the downstream of constructed reclaimed water treatment ponds.

Turbidity Poor catchment management of reclaimed water systems can exaggerate the turbidity of water (i.e. the absence or very low density of submerged vegetation). In highly turbid water, the light penetration is reduced, which can affect photosynthesis of plants, breakdown of contaminants by photolysis and cause temperature stratification (i.e. reduced solar heating of the water surface). Many studies suggest that turbidity could be one of the main water quality parameters influencing aquatic macroinvertebrate diversity (20, 21). However, there are very few studies on the likely impacts on the tolerance of certain macroinvertebrate groups from which base predictions of likely effects can be drawn. This could be caused by highly variable properties of suspended particles, and thus without clear definition of particulate matter, it is very difficult to conduct comparative toxicity studies. The only study we found in the literature on the effect of turbidity used the survival rate of three macroinvertebrate species and results are summarized in Table 2 (22). It appears that except for Procloeon in very high turbid conditions (i.e. NTU 1000), increased turbidity has no pronounced effect on all macroinvertebrate groups investigated. Another complication of increased suspended solids level is to cause physical effects through abrasion and smothering, which can affect movement, feeding, habitat and reproduction of some macroinvertebrates. One particular group of macroinvertebrates that are sensitive to changes of turbidity are filter feeders (e.g. freshwater bivalves, a class of Mollusca), as the suspended solids can also clog respiratory surfaces or interfere with feeding appendages. Aldridge et al. 360 In Water Challenges and Solutions on a Global Scale; Loganathan, et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

(1987) found that bivalve spent more energy to collect food and received reduced nutritional value with simulated increased turbidity in lab experiments (23).


Table 2. Effects of Suspended Particles on Survival of the Baetid Mayfly Procloeonsp, Amphipod H. Azteca, and Midge C. dilutus in 96-h Experiments (Results from Three Experiments for Each Species) Nominal turbidity (NTUs)

Mean Procloeon survival (%)

Mean Hyalella survival (%)

Mean Chironomus survival (%)

0

73.3

90.3

94.7

250

73.3

92.7

95.3

500

74.3

92.7

96.3

1000

86.7

98.0

94.3

Salinity Most reclaimed water (both treated waste and urban stormwater) contains elevated levels of salt, which sometimes can be 1.5-2 times higher than that of tap water. Aquatic macroinvertebrate communities have varying tolerances to different levels of salinities, and macroinvertebrate salinity sensitivity does not correspond well with taxonomic groups. For example, different genera and species within the same family in some instances may have a markedly different sensitivity to salinity (24). In general, Hart et al indicated that adverse effects of salinity begin at the level of 1.47 mS/cm (25), while James et al concluded that salinity of more than 3 mS/cm would produce lethal effects for most of macroinverbrate community (26). For the class of insecta, sensitive orders such as Ephemeroptera (i.e. mayflies), Plecoptera (i.e. stoneflies) and Trichoptera (i.e. caddisflies) tend to show greatest abundance at sites with lowest salinity, while Hemiptera, Diptera (e.g. the larvae of dragonflies and damselflies) and Coleoptera (e.g. Georissus sp.) can be quite tolerant of salinity increase (27). Studies in the literature have shown that crustaceans and molluscs aregenerally more tolerant to rising salinity than aquatic insects. One more inclusive study investigated 31 species in 6 classes of macroinvertebrates in Lijiang river, China and found that the distribution of Mollusca Gastropod species was positively correlated to salinity in the lower conductivity range (0.14 to 0.25 mS/cm) (28). Authors inferred that some gastropod species that need a certain amount of Ca2+ to form shells would prefer higher salinity (29). Only exception in this study was C. cahayensis, which lives in muddy beds and may absorb Ca2+ from silt sediment other than the water body. Similar effect of salinity on macroinvertebrate communities was also observed in the study of the northwest Mississippi River (30). However, very saline and hard water can hardly be favorable for the development of any gastropod populations (31). 361 In Water Challenges and Solutions on a Global Scale; Loganathan, et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Table 3. Summary of Acute 72-hour Salinity Tolerance of Selected Macroinvertebrates to Increased Salinity (32) LC50*

95% CI

centroptilum

5.5

0.76 9.8

Genus 1

NA

6.2

3.7–3.9

Chironomidae

NA

NA

10

6.8–15.0

Gastropoda

Physidae

Physa

acuta

14

13–15

Trichoptera

Ecnomidae

Ecnomus

NA

16

9–28

Hemiptera

Corixidae

Micronecta

annae

17

16–29

Plecoptera

Gripopterygidae

Dinotoperla

twaitesi

18

15–24

Trichoptera

Leptoceridae

Triplectides

australicus

22

19–24

Trichoptera

Calamoceratidae

Anisocentropus

NA

23

19–26

Trichoptera

Leptoceridae

Notilina

spira

25

22–29

Decapoda

Atyidae

Paratya

australiaiensis

38

34–42

Amphipoda

Ceinidae

Austrochiltonia

NA

52

47–59


Order

*

Family

Genus

Species

Ephemeroptera Baetidae

Cloeon

Ephemeroptera Baetidae Diptera

LC50 values are all in mS/cm, NA=Not Available, CI=Confidence Interval.

In literature, the acute tolerance of different macroinverbrate groups to salinity level has been well studied (see Table 3) (28). There are two likely mechanisms that can cause adverse effects on different aquatic macroinvertebrate communities. The first one is related to osmosis and more precisely the ability of a certain macroinvertebrate species to regulate its internal ionic composition against an external gradient. For example,species with the higher internal ionic concentration, such as shrimp, more likely possess higher salinity tolerance. Another important factor could be reduced ability of saline water to hold DO than fresh water, which usually results in changes of macroinvertebrate distribution in aquatic systems (33).

Nutrients Many of the substances present in water are essential to the functioning of aquatic ecosystems under normal conditions, but they can have adverse effects in excessive concentrations. The most common are nutrients (i.e. total nitrogen and phosphorus), which are essential for the growth of algae and other aquatic plants. However, high levels of one or both in reclaimed water received 362 In Water Challenges and Solutions on a Global Scale; Loganathan, et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.


from agricultural and urban sources can lead to excessive plant growth or algal blooms along with other ecological consequences. For example, plant respiration can result in the decrease of DO at night; the death and decay of algae can produce toxins and stagnant conditions. While many macroinvertebrates do not respond directly to nutrient enrichment, they have been shown to respond to nutrient-induced water quality deterioration, such as decrease of DO. In these conditions, macroinvertebrate community diversity and abundance is usually reduced, especially for sensitive species of Plecoptera and Trichoptera. For sensitive species of Ephemeropterans, their tolerance to DO varies widely, with survival response LC50 values ranging from 5.3 mg/L for Rhithrogenairidina to 0.03 mg/L for Ephemera vulgate (34, 35). Many species of molluscs appear to be sensitive to DO levels below 6.0 mg/L (36). Distribution of the species of Mollusca Lamellibranchiata (e.g. C. Fluminea) was reported to be negatively related to the total nitrogen level, while C. Cahayensis of mollusca Gastropoda was reported to prefer a more enriched environment where the total nitrogen level are within the range of 0-2.29 mg/L (28). In aquatic systems with low DO level, there is generally an increase in the abundance of a few opportunistic species, which are able to take advantage of the altered conditions and exploit the excess of food supply. For instance, species of Hemiptera (e.g. Neogerris, Palmcorixa, Ramphocorixa, and hesperocorixa), Diptera (e.g. C. plumosus-a larval Chironomidae), Coleoptera and Oligochaeta (e.g. Tubifextubifex,Limnodrilushoffmeisteri and L. udekemianus) can dominate the macroinvertebrate assemblage in high total nitrogen and phosphorous environments and are very tolerant to low levels of dissolved oxygen. Species of Hirudinea and Decapoda (e.g. Hyallelaazteca, Gammaruslacustris) can reportedly tolerate DO levels below 1.0 mg/L (37–40). It has been reported that upland Ephemeroptera and lowland Chironomidae were the most and least sensitive taxa, respectively, in Australia (41).

Toxic Metals Heavy metals such as As, Cd, Cu, Cr, Ni, Pb, and Zn are frequently detected in most reclaimed wastewater sources, however, their presence is largely negligible and the concentrations are comparable to the levels found in fresh water. This is because conventional wastewater treatment process can remove them from the water stream effectively and they tend to concentrate in the sludge fraction. On the other hand, metal concentrations in urban reclaimed stormwater are typically much higher, as a result of untreated run-off from roads and car parks, deterioration of building surfaces (e.g. roofs), air conditioning coolants, pesticides, batteries and electroplating. However, their concentration in receiving reclaimed water systems could be much less than in undiluted stormwater. Ecotoxicological investigations of water contaminants can be classified according to their level of complexity into experimental micro- and mesocosm studies in laboratory and field systems, artificial contamination of streams, in situ bioassays and field studies about the effects of the ‘natural’ contamination (42). Among aquatic macroinvertebrates, metal toxicity studies on aquatic insects have 363 In Water Challenges and Solutions on a Global Scale; Loganathan, et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.


received most attention. Many laboratory studies showed that aquatic insects are relatively insensitive to acute metal exposure (especially Cd and Cu) (43), using species mainly from 4 orders-Diptera, Ephemeroptera, Plecoptera, and Trichoptera. However, some field studies suggest particular aquatic insects (e.g. some species of Ephemeroptera) are among the most vulnerable to metal exposure in many aquatic systems (44). This discrepancy in literature findings can be attributed to a number of factors. 1) There is usually a presence of multiple metals in field studies, while most laboratory studies are restricted to single insect species exposed to a single metal toxicant. 2) There is inconsistency of sample size, exposure duration and endpoint across the studies. 3) The toxic significance of metal concentrations on aquatic macroinvertebrates is often difficult to interpret, because the bio-available fraction is usually unquantified in the study (45). 4) There are very limited studies in the literature on the relative importance of waterborne versus dietary pathway of metals exposure to aquatic insects. Among all the toxic metals detected in water, there are only a few detailed studies on the pathways of Cd exposure, which suggest dietary pathway is dominant form for most aquatic insects (46–49), and one study on the pathway of Zn exposure to caddisfly larvae (50). Given the limited data available, there is a clear need for developing a more mechanistic understanding of aquatic insect sensitivity to toxic metals especially in long–term laboratory and field studies.

Organic Pollutants Organic pollutantscan enter reclaimed water from various urban, industrial and agricultural sources, including pharmaceuticals, pesticides, personal care products (e.g. surfactants), petrochemicals (e.g. polycyclic aromatic hydrocarbon), and disinfection by-products (DBPs) (51, 52). In most cases, however, concentrations of organic toxicants detected in reclaimed water vary considerably due to their different degree of removal in the wastewater treatment plant. The presence of these organic toxicants can generally result in the decrease of overall biodiversity, in the simplification of macro-invertebrate structures, and in the distortion of functional feeding group composition. The effect to the macroinvertebrate communities may be either short-term (acute), if the pollutant exists in the water at high enough concentrations, or long-term (chronic) where toxins can accumulate and become concentrated in food chains. Symptoms associated with organic toxicants include decreased reproduction, impaired behavioural responses, disease or eventually death. The response of different macroinvertebrates to organic pollutants can vary enormously. For example, most species of mayfly (i.e. Ephemeroptera) nymph do not respond well to sediment or organic pollutants, but some are quite tolerant. Short-term toxicity tests on aquatic macroinvertebrates often produce mixed results, although longer-term in-situ toxicity tests have more consistently demonstrated the potential toxic effects of urban stormwater runoff (53, 54). It should also be noted that toxic effects of aquatic contaminants can be measured and assessed in two ways by concentrations and loads. Concentrations are measures of how much of a contaminant is in a fixed volume of water (units in mass/volume, e.g. mg/L). Loads are measures 364 In Water Challenges and Solutions on a Global Scale; Loganathan, et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.


(always estimated) of how much contaminant is transported by a stream over a period of time (units in mass/time, e.g. kg/yr). While laboratory studies tend to focus on the effects of each toxicant based on their concentrations, field studies can be conducted in both ways. Until now, studies on the toxic effects of insecticides and herbicides have received most attention (55–60). For example, several studies suggest pesticide residual in the aquatic environment could be the number 1 stressor to the aquatic macroinvertebrate community, especially when the water body received agricultural run-off (22). Common classes of insecticides include organochlorides, organophosphates, carbamates, pyrethroids, neonicotinoids, ryanoids, etc. Dassanayake et al (2003) studied three different types of pesticide (i.e. Atrazine a triazine herbicide, Molinate a thiocarbamate pesticide and Chlorpyrifos an organophosphate insecticide) using Daphnia carinata (Crustacea: Cladocera) (61). The toxicity of pesticide was found to increase with salinity, except for Chlorpyrifos. Monserrat et al investigated the toxic effect of another organophosphate pesticide (i.e. parathion) on Chasmagnathusgranulata (Decapoda, Grapsidae) and confirmed that their sensitivity to the pesticide tended to elevate with the salinity (62). A field study in a constructed wetland at the University of Mississippi Field Station investigated toxic effects of two pesticides (i.e. organophosphate Diazinon and pyrethroidpermethrin) on macroinvertebrate community using Hyalella Azteca as indicator and found that the vegetation are effective in mitigating the toxicity during a short-term test (63). Studies on toxic effects of neonicotinoid insecticides such as thiacloprid and imidacloprid were also reported (64, 65). The only study we found in literature using a multi-species assemblage consisting of Gammaruspulex (amphipod), Leuctra nigra (stonefly), Heptageniasulphurea (mayfly) and Ancylusfluviatilis (gastropod) was to investigate the effect of lambda-cyhalothrin (i.e. a pyrethroid insecticide) and carried out in outdoor experimental stream channels (66). Pharmaceuticals is another major group of organic contaminants increasingly detected in reclaimed water via biomedical, veterinary medicine, agricultural and industrial routes, including β-blockers (e.g. Metoprolol up to 1.54 mg/L) and beta-sympathomimetics, analgesic and anti-inflammatory drugs (e.g. Diclofenac up to 1.2 mg/L), endocrine disrupting compounds (EDCs) (e.g. 17b-estradiol up to 0.013 mg/L) andalso antibiotics (e.g. Erythromycin up to 1.7 mg/L) as well as lipid lowering agents(e.g. Clofibrinic acid up to 0.2 mg/L) and psychiatric drugs (e.g. Carbamazepineup to 2.1 mg/L) (67, 68). So far, the majority of studies carried out were limited to controlled laboratory environments and using Daphnia magna, commonly known as the water flea, as a model indicator due to its relatively short life span (7-8 weeks). A recent review by Dang et al compared all the studies in the literature and reported results on the effects of 86 known EDCs on Daphnia magna (69), while they indicated that the effects were no conclusively EDC-related. Other types of pharmaceuticals reported in laboratory studies include anti-inflammatory drugs (70, 71), psychiatric drugs (72), β-blockers (73, 74), and anti-microbial drugs (75). The only field study we found in literature was conducted in the Llobregat river, Spain, and investigated toxic effects of 64 different types of pharmaceuticals on caddisfly larvae (H. exocellata) and amphipod (Echinogammaruslongisetosus) (76). 365 In Water Challenges and Solutions on a Global Scale; Loganathan, et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.


In the reclaimed water systems, polycyclic aromatic hydrocarbon (PAHs) contamination from petroleum runoff, industrial processes, and petroleum spills, can be phototoxic to many aquatic macroinvertebrates. The study by Hatch et al demonstrated the acute photo-induced toxicity of anthracene and fluoranthene on the survival of the midge Chironomustentans and the freshwater amphipod Hyalella Azteca (77). Besides the dose, controlled laboratory experimental studies confirmed that PAH phototoxic effect is also dependent on light intensity, photoperiod, and spectrum (78, 79). In the field, increased turbidity and dissolved humic substances has reportedly shown attenuating effects on acute photo-induced toxicity in aquatic biota by reducing their exposure to UV (80, 81). Surfactants in reclaimed water is another group of concern due to the extensive usage of detergents, shampoo and other household cleaners (82), and there are several studies on the toxic effects of different surfactants (e.g. alkylphenols, anionic sulphonates) using Daphnia and other cladocera (83–86). Their toxic levels, as in macroinvertebrate survival rate, are generally reported to be much lower than those of pesticides and pharmaceuticals. Perfluoro-chemicals, as an emerging group of organic contaminants, and their effects on aquatic macroinvertebrates have also been reported, though very limited, such as perfluorooctanesulfonate (PFOS) and perfluorooctanoic acid (PFOA) (87, 88). Given the fact that many organic contaminants are present at the same time in reclaimed water, it is difficult to predict what the likely combination effects of different organics on macroinvertebrate communities will be, and thus serious consideration should be given to the effect of their interactions where these contaminants are found.

Conclusions Surface-flow type constructed wetlands are being commonly prescribed by local authorities for stormwater and wastewater treatment in many parts of Australia, and elsewhere. These wetlands, though ‘artificial’, have a potential to develop and become complex ecosystems with a variety of fauna and flora, among which macroinvertebrate communities could play an important role in the ecological functioning of healthy aquatic ecosystems as they process detritus and algae, and also provide food for other aquatic animals. Major health concerns from man-made reclaimed water systems could be mosquito-borne disease, including Ross River virus, malaria, West Nile virus, filariasis, yellow fever, although only a small proportion of mosquito species are known vectors for human disease. Water quality conditions of constructed wetlands can impact biodiversity and abundance of different local macroinvertebrate communities greatly. For example, stormwater, not heavily polluted with organic material, may create less of a mosquito problem than sewage or wastewater with higher levels of nutrients to support greater production of vegetation and mosquitoes; Stormwater run-off from tarred roads or other surfaces with chemical deposits, may be potentially lethal to mosquito larvae (and also their predators). Therefore, the potential impact of different reclaimed water quality parameters on aquatic 366 In Water Challenges and Solutions on a Global Scale; Loganathan, et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.


macroinvertebrate assemblage has been reviewed systematically in this chapter and some of major findings and gaps in literature are summarized as follows: 1) At the moment, there are no guideline values of each water quality parameter in the literature that are specifically protective of aquatic macroinvertebrate assemblage. Some studies have proposed water quality indices to evaluate the habitat suitability or biodiversity (89). 2) The effects of altered water quality conditions on macroinvertebrate assemblage are complex and the impacts of different contaminants are interrelated, which may have additive effects (the total effect equals the sum of the individual effects) or even synergistic effects (the total effect exceeds the sum of the individual effects). For instance, increased concentrations of suspended particulate matter can reduce the toxicity of some contaminants probably due to its certain adsorptive properties. A comprehensive review by Hall and Anderson (1995) summarized possible interactive effects between different aqueous contaminants and salinity, including heavy metals, PAHs, organophosphates, etc (90). Water with higher salinity was reported to suppress the toxic effect of heavy metals including cadmium, chromium, copper, mercury, nickel and zinc, due to the reduced bioavailability of the free metal ion (which is the more toxic form) at higher salinities (90, 91). However, interactive effects of organic contaminants with salinity was mixed and in many cases inconclusive (90). Changes in pH or DO level can also affect other contaminants in the water (i.e. heavy metal and phosphorous). For example, cadmium stays in a solid form in the presence of oxygen and sinks to the bottom of the water, while the anoxic (without oxygen) condition may dissolve more cadmium into the water. Therefore the reliance on chronic exposure guidelines or trigger values of single contaminants to assess the toxicity of reclaimed water is likely to result in an underestimation of ecosystem impacts. 3) Most of studies conducted toxicity tests of individual contaminants in the absence of flow-related stresses. The toxic effects of urban stormwater on macroinvertebrates are likely to be greater when associated with flowrelated disturbances following storm events.Furthermore, contaminants may alter the effects of flow disturbances and vice-versa: for instance, in a high-flow event a rock-clinging species stressed by the presence of a toxicant may be more likely to be dislodged (thereby increasing the risk of death) than an unstressed one. This chapter underscores the importance of promoting biodiversity through the maintenance of healthy wetlands (i.e. water quality monitoring and management aspects) and sustainable ecosystem services which could benefit human health. Healthy wetlands should be characterized by intact wetland communities with increased biodiversity and trophic structure that tend to minimize dominance and production of vector mosquito species, reservoir host species and minimize risk of disease to surrounding human and animal populations. Therefore, it would be interesting to know if there is a competing 367 In Water Challenges and Solutions on a Global Scale; Loganathan, et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.


factor in each water quality to promote good species (e.g. predator species) and reduce nuisance species (e.g. mosquitos and biting midges) at the same time.Anywhere where such a difference exists in a pair of organisms there may be an opportunity to manipulate conditions to favour one at the expense of the other, thereby minimising the cause of the nuisance or disease vectoring species.Research into the factors affecting biodiversity in these systems should permit guidelines to be formulated that will inform the future design, construction and operation of such systems to maximize environmental and health benefits; thus ensuring the “friend” scenario strongly outweighs the “foe” situation.

Acknowledgments Authors are grateful to University of South Australia for providing Grant Application Development Fund support for the preparation of this book chapter.

References 1. 2.

3.

4.

5.

6.

7.

8.

9.

Wheeler, S. A. Insights, lessons and benefits from improved regional water security and integration in Australia. Water Resour. Econ. 2014, 8, 57–78. Song, M. Y.; Leprieur, F.; Thomas, A.; Lek-Ang, S.; Chon, T. S.; Lek, S. Impact of agricultural land use on aquatic insect assemblages in the Garonne river catchment (SW France). Aquat. Ecol. 2009, 43, 999–1009. Porinchu, D.; MacDonald, G. The use and application of freshwater midges (Chironomidae, Insecta, Diptera) in geographical research. Prog. Phys. Geogr. 2003, 27, 378–422. Vondracek, B.; Blann, K.; Cox, C.; Nerbonne, J.; Munford, K.; Berbonne, B. Land use, spatial scale and stream systems: lessons from an agricultural region. Ecol. Appl. 2005, 36, 775–791. Zou, L.; Zhang, W., Key contaminants for water purification in view of remediation by nanomaterials. In Applications of nanomaterials for water quality; Van der Bruggen, B., Ed.; Future Science Ltd.: London, 2013; pp 30−50. Tennessen, K. J., Production and suppression of mosquitoes in constructed wetlands. In Constructed Wetlands for Water Quality Improvement; Moshiri, G. A., Eds.; Lewis Publishers (CRC Press): Chelsea, MI, 1993; pp 591–601. Walton, W. E.; Workman, P. D. Effect of mash design on the abundance of mosquitoes in experimental constructed wetlands in Southern California. J. Am. Mosquito. Control Assoc 1998, 14, 95–107. Medlock, J. M.; Vaux, A. G. C. Assessing the possible implications of wetland expansion and management on mosquitoes in Britain. J. Europe. Mosquito. Control Assoc. 2011, 29, 38–65. Mullen, G. Biting midges (Ceratopogonidae). In Medical and Veterinary Entomology; Mullen, G., Durden, L., Eds.; Elsevier Science: San Diego, CA, 2002; pp 163−183. 368 In Water Challenges and Solutions on a Global Scale; Loganathan, et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.


10. Whelan, P. I. Mosquitoes of Public Health Importance in the Northern Territory and their Control; Medical Entomology Branch, Department of Health: Darwin, NT, 1984; p 79. 11. Russell, R. C. Mosquitoes and Mosquito-borne Disease in Southeastern Australia, revised ed.; Department of Medical Entomology, Westmead Hospital: Westmead, NSW, 1993; p 320. 12. Kay, B. H.; Standfast, H. A. Ecology of arboviruses and their vectors in Australia. Curr. Top. Vector Res. 1987, 3, 1–36. 13. Russell, R. C. Arboviruses and their vectors in Australia: an update on the ecology and epidemiology of some mosquito-borne arboviruses. Rev. Med. Vet. Entomol. 1995, 83, 141–158. 14. Service, M. W. Mosquito Ecology: Field Sampling Methods, 2nd ed.; Elsevier: Amsterdam, 1993. 15. Sanchez, L.; Cortinas, J.; Pelaez, O.; Gutierrez, H.; Concepción, D.; Van der Stuyft, P. Breteau Index threshold levels indicating risk for dengue transmission in areas with low Aedes infestation. Trop. Med. Int. Health. 2010, 15, 173–175. 16. Walsh, C. J.; Sharpe, A. K.; Breen, P. F.; Sonneman, J. A. Effects of urbanization on streams of the Melbourne region, Victoria, Australia. I. Benthic macroinvertebrate communities. Freshwater Biol. 2001, 46, 535–551. 17. Hatt, B. E.; Fletcher, T. D.; Walsh, C. J.; Taylor, S. L. The influence of urban density and drainage infrastructure on the concentrations and loads of pollutants in small streams. Environ. Manage. 2004, 34, 112–124. 18. Lessard, J. L.; Hayes, D. B. Effects of elevated water temperature on fish and macroinvertebrate communities below small dams. River Res. Appl. 2003, 19, 721–732. 19. Jacobsen, D.; Schultz, R.; Encalada, A. Structure and diversity of stream invertebrate assemblages: the influence of temperature with altitude and latitude. Freshwater Biol. 1997, 38, 247–261. 20. Dlamini, V.; Hoko, Z.; Murwira, A.; Magaggula, C. Response of awuatic macro-invertebrate diversity to environmental factors along the lower Komati river in Swaziland. Phys. Chem. Earth 2010, 35, 665–671. 21. Van de Meutter, F.; Stoks, R.; De Meester, L. The effect of turbidity state and microhabitat on macroinvertebrate assemblages: apilotstiudy of xsix shallow lakes. Hydrobiologia. 2005, 542, 379–390. 22. Anderson, B. S.; Phillips, B. M.; Hunt, J. W.; Connor, V.; Richard, N.; Tjeerdema, R. S. Identifying primary stressors impacting macroinvertebrates in the Salina River (California, USA): Relative effects of pesticides and suspended particles. Environ. Poll. 2006, 141, 402–408. 23. Aldridge, D. W.; Payne, B. S.; Miller, A. C. The effects of intermittent exposure to suspended solids and turbulence on three species of freshwater mussels. Environ. Poll. 1987, 45, 17–28. 24. Clunie, P.; Ryan, T.; James, K.; Cant, B. Implications for Rivers from Salinity Hazards: Scoping Study; Report to the Murray Darling Basin Commission; Department of Natural Resources and Environment: Melbourne, Victoria, 2002. 369 In Water Challenges and Solutions on a Global Scale; Loganathan, et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.


25. Hart, B. T.; Bailey, P.; Edwards, R.; Hortle, K.; James, K.; McMahon, A.; Meredith, C.; Swadling, K. A review of the salt sensitivity of the Australian freshwater biota. Hydrobiologia 1991, 210, 105–144. 26. James, K. R.; Cant, B.; Ryan, T. Responses of freshwater biota to rising salinity levels and implications for saline water management: a review. Aust. J. Bot. 2003, 51, 703–713. 27. Arimoro, F.; Ikomi, R.; Ajuzieogu, I.; Nwadukwe, F. Temporal and spatial variability in macroinvertebrate community structure in relation to environmental variables in Ajijiguan Creek, Niger Delta, Nigeria. Afr. J. Aquat. Sci. 2011, 36, 57–66. 28. Chen, Q.; Yang, Q.; Li, R.; Ma, J. Spring micro-distribution of macroinvertebrate in relation to hydro-environmental factors in the Lijiang Rive, China. J. Hydro-environ. Res. 2013, 7, 103–112. 29. Welch, E. B.; Jacoby, H. M.; May, C. W. Stream quality. In River Ecology and Management; Naiman, R. J., Bilby, R. E., Ed.; Springer: New York, 2001; pp 69−94. 30. Maul, J. D.; Farris, J. L.; Milam, C. D.; Cooper, C. M.; Testa, S.; Feldman, D. L. The influence of stream habitat and water quality on macroinvertebrate communities in degraded streams of northwest Mississippi. Hydrobiologia 2004, 518, 79–94. 31. Ciocco, N.; Scheibler, E. Malacofauna of the littoral benthos of saline lake in southern Mendoza, Argentina. Fundam. Appl. Limnol. 2008, 172, 87–98. 32. Kefford, B. J.; Paradise, T.; Papas, P. J.; Fields, E.; Nugegoda, D. Assessment of a System to Predict the Loss of Aquatic Biodiversity from Changes in Salinity; Report to Land and Water Australia, Project No: VCE 17, Perth, Australia, 2003. 33. Kay, W. R.; Halsem, S. A.; Scanlon, I.; Smith, M. J. Distribution and environmental tolerance of aquatic macroinvertebrate families in the agricultural zone of south-western Australia. J. N. Am. Benthol. Soc. 2001, 20, 132–199. 34. Jacob, U.; Walther, H. Aquatic insect larvae as indicators of limiting minimal contents of dissolved oxygen. Aquat. Insect 1981, 3, 219–224. 35. Jacob, U.; Walther, H.; Klenke, R. Aquatic insect larvae as indicators of limiting minimal contents of dissolved oxygen—Part II. Aquat. Insect 1984, 6, 185–190. 36. Fuller, S. L. H. Clams and mussels (Mollusca: Bivalvia). In Pollution ecology of freshwater invertebrates; Hart, C. W., Jr., Fuller, S. L. H., Eds.; Academic Press, Inc.: New York, 1974; pp 215–274. 37. Hobbs, H. H.; Hall, E. T. Crayfishes (Decapoda: Astacidae). In Pollution ecology of freshwater invertebrates; Hart, C. W., Jr., Fuller, S. L. H., Eds.; Academic Press, Inc.: New York, 1974; pp 195–214. 38. Sawyer, R. T. Leeches (Annelida: Hirudinea). In Pollution Ecology of Freshwater Invertebrates; Hart, C. W., Jr., Fuller, S. L. H., Eds.; Academic Press, Inc.: New York, 1974, pp 82–142. 39. Nebecker, A. V. Effect of low oxygen concentration on survival and emergence of aquatic insects. Trans. Am. Fish. Soc. 1972, 4, 675–679. 370 In Water Challenges and Solutions on a Global Scale; Loganathan, et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.


40. Nebecker, A. V.; Onjukka, S. T.; Stevens, D. G.; Chapman, G. A.; Dominguez, S. E. Effects of low dissolved oxygen on survival, growth, and reproduction of Daphnia, Hyalella, and Gammarus. Environ. Toxicol. Chem. 1992, 11, 373–379. 41. Connolly, N. M.; Crossland, M. R.; Pearson, R. G. Effect of low dissolved oxygen on survival, emergence, and drift of tropical stream macroinvertebrates. J. N. Am. Benthol. Soc. 2004, 23, 251–270. 42. Buikama, A. L. Jr.; VoshellJr.J. R.. Toxicity studies using freshwater benthic macroinvertebrates. In Freshwater Biomonitoring and Benthic Macroinvertebrates; Rosenberg, D. M., Resh, V. H., Eds.; Chapman and Hall: New York, 1993: pp 344−398. 43. Brix, K. V.; DeForest, D. K.; Burger, M.; Adams, W. J. Assessing the relative sensitivity of aquatic organisms to divalent metals and their representation in toxicity data sets compared to natural aquatic communities. Hum. Ecol. Risk. Assess. 2005, 11, 1139–1156. 44. Clements, W. H.; Carlisle, D. M.; Lazorchak, J. M.; Johnson, P. C. Heavy metals structure benthiccommunities in Colorado mountain streams. Ecol. Appl. 2000, 10, 626–38. 45. Welch, E. B.; Jacoby, J. M.; May, C. W. Stream quality. In River Ecology and Management. Lessons from the Pacific Coastal Ecoregion; Naiman R. J., Bilby, R. E., Eds.; Springer: New York, 1998; pp 69–94. 46. Martin, C. A.; Luoma, S. N.; Cain, D. J.; Buchwalter, D. B. Cadmium ecophysiology in seven stonefly (Plecoptera) species: delineating sources and estimating susceptibility. Environ. Sci. Technol. 2007, 41, 7171–7177. 47. Munger, C.; Hare, L. Relative importance of water and food as cadmium sources to an aquatic insect (Chaoboruspunctipennis): implications for predicting Cd bioaccumulation in nature. Environ. Sci. Technol. 1997, 31, 891–895. 48. Roy, I.; Hare, L. Relative importance of water and food as cadmium sources to the predatory insect Sialiavelata (Megaloptera). Can J. Fish Aquat. Sci. 1999, 556, 1143–149. 49. Xie, L; Funk, D. H.; Buchwalter, D. B. Tropic transfer of Cd from natural periphyton to the grazing mayfly Centroptilumtriangulifer in a life cycle test. Environ. Pollut. 2010, 158, 272–277. 50. Timmermans, K. R.; Spijkerman, E.; Tonkes, M.; Govers, H. Cadmium and zinc uptake by two species of aquatic invertebrate predatros from dietary and aqueous soures. Can. J. Fish Aquat. Sci. 1992, 49, 655–62. 51. Kimbrough, R. A.; Litke, D. W. Pesticides in streams draining agricultural and urban areas in Colorado. Environ. Sci. Technol. 1996, 30, 908–916. 52. Schroeter, H. O. Toxic contaminant loadings from municipal sources in Ontario areas of concern. Water Qual. Res. J. Can. 1997, 32, 7–22. 53. Pesacreta, G. J. Response of the stonefly Pteronarcysdorsatain enclosures from an urban North Carolina stream. Bull. Environ. Contam. Toxicol. 1997, 59, 948–955. 54. Burton, G. A.; Pitt, R. E.; Clark, S. The role of traditional and novel toxicity test methods in assessing stormwater and sediment contamination. Crit. Rev. Environ. Sci. Technol. 2000, 30, 413–447. 371 In Water Challenges and Solutions on a Global Scale; Loganathan, et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.


55. Schulz, R.; Mathias, L. A field study of the effects of agriculturally derived insecticide input on stream macroinvertebrate dynamics. Aquat. Toxicol. 1999, 46, 155–176. 56. VanWijngaarden, R. P. A.; Brock, T. C. M.; van Den Brink, P. J. Threshold levels for effects of insecticides in freshwater ecosystems: a review. Ecotoxicology. 2005, 14, 355–80. 57. Schulz, R.; Moore, M. T.; Bennett, E. R.; Milam, C. D.; Bouldin, J. L.; Farris, J. L.; Smith, S., Jr.; Cooper, C. M. Acute toxicity of methylparathion in wetland mesocosms: assessing the influence of aquatic plants using laboratory testing with Hyalellaazteca. Arch. Environ. Contam. Toxicol. 2003, 45, 331–336. 58. Smith, S.; Lizotte, R. E. Influence of selected water quality characteristics on the toxicity of k-cyhalothrin and c-cyhalothrin to Hyalellaazteca. Bull. Environ. Contam. Toxicol. 2007, 79, 548–551. 59. Bouldin, J. L.; Farris, J. L.; Moore, M. T.; Smith, S.; Cooper, C. M. Assessment of diazinon toxicity in sediment and water of constructed wetlands using deployed Corbiculafluminea and laboratory testing. Arch. Environ. Contam. Toxicol. 2007, 53, 174–182. 60. Relyea, R. A. The impact of insecticides and herbicides on the biodiversity and productivity of aquatic communities. Ecol. Appl. 2005, 15, 618–627. 61. Dassanayake, H.; Warne, M. S. J.; Lim, R. P. Interactive effects of salinity on the toxicities of Atrazine, Molinate, and Chlorpyrifos to the Cladoceran Daphnia carinata. SETAC Asia Pacific, Australasian Society for Ecotoxicology; 2003; p 50. 62. Monserrat, J. E.; Rodrígues, E. M.; Lombardo, R. J. Effects of salinity on the toxicity of parathion to the estuarine crab Chasmagnathusgranulata (Decapoda, Grapsidae). Bull. Environ. Contam. Toxicol. 1991, 46, 569–575. 63. Lizotte, R., Jr.; Moore, M.; Locke, M.; Kroger, R. Role of vegetation in a constructed wetland on nutrient-pesticide mixture toxicity to Hyalella Azteca. Arch. Environ. Contam. Toxicol. 2011, 60, 261–271. 64. Sánchez-Bayo, F.; Goka, K. Influence of light in acute toxicity bioassays of imidacloprid and zinc pyrithione to zooplankton crustaceans. Aquat. Toxicol. 2006, 78, 262–271. 65. Beketov, M. A.; Schäfer, R. B.; Marwitz, A.; Paschke, A.; Liess, M. Longterm stream invertebrate community alterations induced by the insecticide thiacloprid: effect concentrations and recovery dynamics. Sci. Total Environ. 2008, 405, 96–108. 66. Rasmussen, J. J.; Friberg, N.; Larsen, S. E. Impact of lambda-cyhalothrin on a macroinvertebrate assemblage in outdoor experimental channels: Implications for ecosystem functioning. Aquat. Toxicol. 2008, 90, 228–234. 67. Fent, K.; Weston, A.; Caminada, D. Ecotoxicology of human pharmaceuticals. Aquat. Toxicol. 2006, 76, 122–159. 68. Rosi-Marshall, E. J.; Royer, T. V. Pharmaceutical compounds and ecosystem function: an emerging research challenge for aquatic ecologists. Ecosystems. 2012, 15, 867–880. 372 In Water Challenges and Solutions on a Global Scale; Loganathan, et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.


69. Dang, Z.; Cheng, Y.; Chen, H.; Cui, Y.; Yin, H.; Traas, T.; Montforts, M.; Vermeire, T. Evaluation of the Daphnia magna reproduction test for detecting endocrine disruptors. Chemosphere. 2012, 88, 514–523. 70. Cleuvers, M. Mixture toxicity of the anti-inflammatroy drugs diclofenac, ibuprofen, naproxen, and acetylsalicylic acid. Ecotox. Environ. Safety. 2004, 59, 309–315. 71. Heckmann, L.; Callaghan, A.; Hooper, H. L.; Connon, R.; Hutchinson, T. H.; Maudnd, S. J.; Sibly, R. M. Chronic toxicity of ibuprofen to Daphnia magna: Effects on life history traits and population dynamics. Toxicol. Lett. 2007, 172, 137–145. 72. Jarivs, A. L.; Bernot, M. J.; Bernot, R. J. Relationshipd between the psychiatric drug carbamazepine and freshwater macroinvertebrate community structure. Sci. Total Environ. 2014, 496, 499–509. 73. Franzellitti, S.; Buratti, S.; Valbonesi, P.; Capuzzo, A.; Fabbri, E. The b-blockerpropranolol affects cAMP-dependent signaling and induces the stress response in Mediterranean mussels, Mytilusgalloprovincialis. Aquat. Toxicol. 2011, 101, 299–308. 74. Cleuvers, M. Initial risk assessment for three found in the aquatic environment. Chemosphere. 2005, 59, 199–205. 75. Anskjær, G. G.; Rendal, C.; Kusk, K. O. Effect of pH on the toxicity and bioconcentration of sulfadiazine on Daphnia magna. Chemosphere. 2013, 91, 1183–1188. 76. Damásio, J.; Barceló, D.; Brix, R.; Gros, C. P.; M.; Petrovic, M.; Sabater, S.; Guasch, H.; de Alda, M. L.; Barata, C. Are pharmaceuticals more harmful than other pollutants to aquatic invertebrate species: A hypothesis tested using multi-biomarker and multi-species responses in field collected and transplanted organisms. Chemosphere. 2001, 85, 1548–1554. 77. Hatch, A. C.; Burton, G. A., Jr. Photo-induced toxicity of PAHs to Hyalella Azteca and Chironomustentans: effects of mixtures and behaviour. Environ. Pollut. 1999, 106, 157–167. 78. Wilcoxen, S. E.; Meier, P. G.; Landrum, P. F. The toxicity of fluoranthene to Hyalella Azteca in sediment and water-only expostures under verying light spectra. Ecotox. Environ. Saf. 2003, 54, 105–117. 79. Monson, P. D.; Ankley, G. T.; Kosian, P. A. Phototoxic response of Lumbriculusvariegatus to sediments contaminated by polycyclic aromatic hydrocarbons. Environ. Toxicol. Chem. 1995, 14, 891–894. 80. Ireland, D. S.; Burton, G. A.; Hess, G. G. In situ evaluations of turbidity and photoinduction of polycyclic aromatic hydrocarbons. Environ. Toxicol. Chem. 1996, 15, 574–581. 81. Oris, J. T.; Hall, A. T.; Tylka, J. D. Humic acids reduce the photo-induced toxicity of anthracene to fish and Daphnia. Environ. Toxicol. Chem. 1990, 9, 575–583. 82. Lewis, M. A. Chronic and sublethal toxicities of surfactants to aquatic animals:a review and risk assessment. Water Res. 1991, 25, 101–113. 83. Dave, G.; Herger, G. Determination of detoxification to Daphnia Magna of four pharmaceuticals and seven surfactants by activated sludge. Chemosphere. 2012, 88, 459–466. 373 In Water Challenges and Solutions on a Global Scale; Loganathan, et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.


84. Hodges, G.; Roberts, D. W.; Marshall, S. J.; Dearden, J. C. The aquatic toxicity of anionic surfactants to Daphnia magna – a comparative QSAR study of linear alkylbenzene sulphonates and ester sulphonates. Chemosphere 2006, 63, 1443–1450. 85. Isidori, M.; Lavorgana, M.; Nardelli, A.; Parella, A. Toxicity on crustaceans and endocrine disrupting activity on Saccaromycescervisiae of eight alkylphenols. Chemosphere. 2006, 64, 135–143. 86. Da Silva Coelho, K.; Rocha, O. Assessment of the potential toxicity of a linear alkylbenzenesulfonate (LAS) to freshwater animal life by means of cladoceran bioassays. Ecotoxicology. 2008, 19, 812–818. 87. Boudreau, T. M.; Sibley, P. K.; Mabury, S. A.; Muir, D. G. C.; Solomon, K. R. Laboratory evaluation of the toxicity of perfluorooctanesulfonate (PFOS) on Selenastrumcapricornutum, Chlorella vulgaris, Lemnagibba, Daphnia magna, and Daphnia pulicaria. Arch. Environ. Contamin. Toxicol. 2003, 44, 307–313. 88. Ji, K.; Kim, Y.; Oh, S.; Ahn, B.; Jo, H.; Choi, K. Toxicity of perfluorooctane sulfonic acid and perfluorooctanoic acid on freshwater macroinvertebrates (Daphnia magna and Moina macrocopa) and fish (Oryziaslatipes). Environ. Toxicol. Chem. 2008, 27, 2159–2168. 89. Duan, X.; Wang, Z.; Xu, M. Effects of fluvial processes and human activities on stream macro-invertebrates. Int. J. Sediment Res. 2011, 26, 416–430. 90. Hall, L. W. J.; Anderson, R. D. The influence of salinity on the toxicity of various classes of chemicals to aquatic biota. Crit. Rev. Toxicol. 1995, 25, 281–341. 91. Voyer, R. A.; McGovern, D. G. Influence of constant fluctuating salinity on responses of Mysidopsisbahiaexposed to cadmium in a life-cycle test. Aquat. Toxicol. 1991, 19, 215–230.

374 In Water Challenges and Solutions on a Global Scale; Loganathan, et al.; ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Reclaimed Water Systems: Biodiversity Friend or ... - ACS Publications

Recommend Documents