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Water Quality Aquatic Life 4.1 Effects Organic carbon plays a central role in lake and stream chemistry because it complexes metals and nutrients, affects pH and alkalinity, and acts as a substrate for microbial production (Dillon and Molot 1997). Dissolved organic carbon (DOC) can protect as well as harm aquatic life depending on concentration, chemical composition and system characteristics. For example, the humic fraction of DOC can form particulates that serve as a food source for filter and deposit feeders, while at the same time increasing exposure of these organisms to higher levels of non-polar toxic chemicals (Björk and Gilek 1996). Petersen and Persson (cited in Kullberg et al. 1993) observed that at pH 5, increasing humus concentrations improved survivorship and growth of Daphnia magna up to 10 mg/L after which survivorship declined. This section focuses on the role of organic carbon in the following key aquatic system processes:
Organic carbon also affects the optical properties of water which can have important consequences for protection of biota from the effects of ultraviolet radiation, depth of the euphotic zone, and algal species composition (see BC Environment 1998a). Because these effects were the focus of the recently developed water quality criteria for true and apparent colour (BC Environment 1998a), they will not be dealt with here.
Numerous studies have shown that the humics and other compounds that comprise dissolved organic carbon in water complex with cadmium (Stephenson and Mackie 1988), aluminum (Simonin et al. 1993), silver (Janes and Playle 1995), zinc (Starodub et al. 1987), lead (Starodub et al. 1987) and copper (Wilson 1972; Nilssen 1982; Winner 1985; Starodub et al. 1987; Meador 1991; Breault et al. 1996), thus reducing the toxicity of these metals (Winner 1985; Starodub et al. 1987; Hutchinson and Sprague 1987; Meador 1991; Welsh et al. 1993; Bresler et al. 1995). Similar relationships have been observed between organic carbon and bioavailability of organics such as PAHs and PCBs (Broman et al. 1996), chlorothalonil (Winkler et al. 1996), and cationic surfactants (Lewis 1992). Humic substances in dissolved organic carbon can bind with phosphorus or form iron-phosphate complexes, thus reducing the available phosphorus (Kullberg et al. 1993). The proportion of metals complexed to humus is dependent on pH, ionic strength, and the relative concentrations of metals and humic material. At low pH, a larger portion of metals is released from the humic acid ligands, thus increasing bioavailability (Starodub et al. 1987; Spry and Wiener 1991; Meador 1991; Welsh et al. 1993). Hutchinson and Sprague (1987) showed the ameliorating effect of natural total organic carbon by exposing flagfish fry to a mixture of aluminum, copper and zinc at pH 5.8 in lake waters with natural TOC ranging from 1.9 to 8.5 mg/L. The LC50 in the highest TOC water was about 7-fold greater than that for the lowest TOC water. The following regression equation explained 84% of the variation in lethality of the metals mixture.
Log10 LC50 = -0.395 + 1.145 - log10TOC
Log10 LC50 = 1.033 + 0.999 - log10DOC
Percent Mortality = 49.7 + 0.01 - CONCDUR - 6.46 - DOC If CONCDUR is held constant, the regression equation indicates that for every 1 mg/L reduction in DOC, mortality would increase 6.46% (until 100% mortality is reached). This study is consistent with the findings of several others (e.g., Effler et al. 1985; Baldigo and Murdoch 1997) that DOC complexes with inorganic Al to create organomonomeric aluminum (Alom) species that are less toxic than Alim to fish (Matuszek and Beggs 1988). Several studies have shown that the relationship between mercury and organic carbon is different from that observed with other metals. Mercury availability, bioaccumulation, and hence toxicity increase as dissolved organic carbon increases in drainage lakes (Nilsson and Håkanson 1992; Driscoll et al. 1994; Mierle and Ingram 1991). Part of the reason for this relationship is that the mercury brought to an aquatic system from the surrounding catchment area is attached to humic substances (Nilsson and Håkanson 1992; Mierle and Ingram 1991). For deep lakes in Sweden, regression analyses indicated that an increase in water colour (a surrogate for dissolved organic carbon) from 10 to 50 mg/L Pt was associated with an increase in mercury tissue levels in fish of between 23.4 and 82.8%, depending on fish species, location, and depth of lake (Nilsson and Håkanson 1992). This suggests that DOC-mediated transport of mercury from the watershed may reduce tissue mercury levels more than the DOC-mediated inhibition of biomethylation which has been observed in several studies (e.g., Barkay et al. 1997). The positive relationship between DOC and tissue mercury levels observed in drainage lakes does not exist for seepage lakes (Grieb et al. 1990). In the case of drainage lakes, much of the mercury is first complexed with humic material and then transported to the aquatic system from the surrounding catchment area (Driscoll et al. 1994; Richardson et al. 1995). Hence, the positive relationship between DOC and tissue mercury levels. In seepage lakes, however, most of the mercury arrives via atmospheric deposition (Grieb et al. 1990). DOC can inhibit bio-methylation of mercury by controlling the substrate for methylation. This can considerably reduce uptake in biota and thus the relationship between DOC and tissue mercury levels in seepage lakes is the reverse of that observed for drainage lakes.
Several studies have noted that brown-water systems with high concentrations of dissolved organic carbon are associated with increased bacterial productivity and support denitrification (Hessen 1985; Tranvik et al. 1991; Weisner et al. 1994). For example, Hessen (1985) observed that the mean annual biomass of bacteria in holarctic oligotrophic lakes was 7.8-12.1 µg C/L in clear lakes, 10.5 µg C/L in slightly coloured lakes and 16.2-44.1 µg C/L in highly coloured humic lakes. Bacterial and zooplankton biomass were also high relative to algal biomass, strongly supporting the possibility that humics are an essential component of the carbon pool in these systems. Similar relationships between bacterial biomass and organic carbon levels have been observed in organically polluted streams (Cazelles et al. 1991) and contaminated ground waters (Harvey and Barber 1992). Increases in bacterial biomass as a result of increases in organic carbon can cause oxygen depletion as has been observed in the central and eastern basins of Lake Erie (Charlton and Rao 1983). The biological functions of many small lakes and headwater streams is controlled by the amount of externally produced allochthonous carbon, rather than the much smaller quantities of autochthonous carbon produced within the water. In such systems, plankton communities consume more carbon than the algae can themselves produce and are therefore reliant on external energy supplies. As a result, changes in input rates of allochthonous carbon would likely have profound effects on primary production and respiration. Indeed, France et al. (1996) observed that removal of riparian vegetation near boreal lakes resulted in reductions of DOC from 17.8 to 0.4 g/m shoreline/year which translated into an annual reduction in organic inputs to the lakes of 8 to 15%. The reductions in organic carbon inputs were associated with a 4-9% decrease in primary production and an 8-17% decline in respiration. Much of the reduction in external DOC input was due to the reduced transport of leaf litter to the lakes (France and Peters 1995), a factor that was not compensated for by increased soil erosion. Lakes and streams affected by clear cutting would likely shift from allotrophy (external energy reliance) to autotrophy (energy self reliance), a change that can have important effects on invertebrate and fish communities (Bilby and Bisson 1992). Dissolved organic carbon in streams complexes with ferrous iron (FeII) and maintains it in this state until it reaches a lake or estuary where it can be utilized by phytoplankton. In the absence of DOC, most iron in flowing streams is oxidized to ferric hydroxide, which phytoplankton cannot assimilate (Kawaguchi et al. 1994). Thus, reductions in DOC from deforestration, agricultural disturbance to the riparian zone, or residential development can have important consequences on watershed productivity, in addition to the localized effects discussed above. Allotrophic aquatic systems receiving large inputs of particulate organic carbon tend to have macroinvertebrate communities with high proportions of shredders and detritivores (Andersson and Danell 1982; Bilby and Bisson 1992; Delong and Brusven 1994). Delong and Brusven (1994) found that in sections of the Lapwai Creek in Idaho with reduced inputs of allochthonous carbon due to agricultural disturbances, shredders were rare and never made up more than 5% of the macroinvertebrate community. In other unaffected streams in the Lapwai Creek drainage basin, shredders were often abundant. Delong and Brusven (1994) also noted that gathering and filtering invertebrates, which are dependent on particulate organic matter, were never as abundant in Lapwai Creek as were grazing organisms utilizing the large biomass of algal matter found in the stream. Abelho and Craça (1996) also noted that not only does quantity of organic carbon inputs affect aquatic communities, so does quality. In their study, they found that replacement of deciduous forests with monocultures of Eucalyptus globulus resulted in reduced numbers and taxa of macroinvertebrates. The reasons for the low quality of eucalyptus detritus are the presence of oil glands and phenols in the leaves, and a superficial wax and a thick cuticle that delay conditioning of the leaves. Many conifer species share these leaf characteristics. Bilby and Bisson (1992) observed that production of coho salmon and shorthead sculpin was elevated in a clear-cut site over that in a nearby old-growth site, despite a five-fold reduction in inputs of allochthonous carbon. In these sites, fish populations appeared to depend on food derived from autotrophic pathways during spring and summer, a hypothesis supported by stomach contents analysis and the similar ratios of autochthonous inputs and fish production between the two streams. In larger lakes and higher order streams, where shading by riparian vegetation has little impact on autotrophs, reductions in inputs of allochthonous organic carbon may lead to reductions in the macro-invertebrate functional groups that depend on these sources as well as the fish species that feed on them (France 1995a).
Declines and losses of fish populations from acidified lakes and rivers have been observed by investigators in Canada, the United States and Scandinavia (Matuszek and Beggs 1988; Schindler 1988; Lien et al. 1996; and numerous others). In Canada, large acid-sensitive areas exist in the mountainous areas of the west, the Yukon, the Northwest Territories, as well as the well-known eastern sector of the country. Acid-sensitive aquatic systems are susceptible to the substantial increases in acidity of catchment soils and water that occurs following deforestation (Hedin et al. 1990). The available evidence suggests that most adult fishes can tolerate pH values of less than 5.5., but that juvenile fishes and many organisms lower in the food web cannot (Schindler 1988; Baldigo and Murdoch 1997). The early disappearance of biota at lower trophic levels may cause starvation to large predatory fish long before the direct toxic effects of the hydrogen ion are evident. The reactive chemistry of dissolved organic matter is of great importance in determining the likely impacts of anthropogenic acid inputs to aquatic systems. Carboxylic acids comprise about 90% of the acid functional groups on humic compounds, and are the most important in the organic carbon pool because they possess reactive sites and contribute aqueous solubility and acidity to organic molecules (Eshleman and Hemond 1985; Kullberg et al. 1993). The presence of acid functional groups provides a pH buffer in aquatic systems, such that between pH 4 and 5, DOC is the principal buffer in surface waters (Kullberg et al. 1993; Roila et al. 1994). Brakke et al. (1988) showed that a large proportion of acidic lakes in the Adirondacks (acid neutralizing capacity or ANC = 0) were clear water lakes with DOC less than 2 mg/L. Hedin et al. (1990) and Roila et al. (1994) estimated that organic anions on humic compounds provide 1.8 to 2.0 meq acid-neutralizing capacity per mg C of DOC. The acid-base properties of DOC also mean that it is a natural background source of acidity, particularly in humus-rich waters, and that recovery of pH will be slowed by DOC buffering (Eshleman and Hemond 1985; Kullberg et al. 1993; Roila et al. 1994). Aluminum is released to lakes and streams during acidification (Hedin et al. 1990). Only inorganic monomeric aluminum (Alim) is highly toxic to fish and present at low pH (Eshleman and Hemond 1985; Brakke et al. 1988; Cosby et al. 1996; Baldigo and Murdoch 1997). Dissolved organic carbon complexes with the Alim produced during acidification and thus if present at sufficiently high levels, DOC can protect fish species from aluminum toxicity (Matuszek and Beggs 1988; Lien et al. 1996; Baldigo and Murdoch 1997). Effler et al. (1985) observed that the removal of aluminum from the photic zone coincided with a depletion of DOC and a decrease in light attenuation; these observations support the notion that DOC can counteract the increased bioavailability of Alim resulting from acidification. The results of the study by Baldigo and Murdoch (1997) on moderately acidic streams (pH 5.2 to 5.6) in the Catskill Mountains, New York suggest that DOC concentrations greater than 2 to 3 mg/L significantly decrease monomeric inorganic aluminum and thus decrease toxicity to brook trout.
Based on a review of the literature and the worldwide web, few, if any, jurisdictions have derived criteria for the protection of marine and freshwater aquatic life, for either dissolved or total dissolved carbon. The CCME (CCREM 1987 and updates) does not have organic carbon guidelines for the protection of freshwater or marine aquatic life.
4.3.1 Total Organic Carbon The 30-day 50th percentile total organic carbon concentration shall be within 20% above or below seasonally-adjusted median background levels as measured historically or at appropriate reference sites. This criterion applies to freshwater aquatic systems only. The 30-day mean calculation should be based on a minimum of five weekly samples taken over a period of 30 days. The appropriate methodology for determining total organic carbon is discussed in section 1.4 and in BC Environment (1994).
The 30-day 50th percentile dissolved organic carbon concentration shall be within 20% above or below seasonally-adjusted median background levels as measured historically or at appropriate reference sites. This criterion applies to freshwater aquatic systems only. The 30-day 50th percentile calculation should be based on a minimum of five weekly samples taken over a period of 30 days. The appropriate methodology for determining dissolved organic carbon is discussed in section 1.4 and in BC Environment (1994).
The review of the literature on effects indicated that changes in concentrations of total and dissolved organic carbon can have a significant impacts on aquatic systems. Reductions in organic carbon, for example, can cause reductions in primary productivity, system metabolism, while increasing susceptibility to toxic metals and acidification. Increases in organic carbon can increase bacterial metabolism to the point of causing anoxia. Changes in the quality of organic carbon entering systems (e.g., allochthonous carbon versus autochthonous carbon) also can have dramatic impacts on the species composition of invertebrate and fish communities. Several of the studies reviewed above indicate that relatively small changes in average organic carbon levels above or below background (about 20%) can have measureable adverse impacts on aquatic systems (e.g., France et al. 1996), especially in areas where metals concentrations are elevated (Hutchinson and Sprague 1987; Simonin et al. 1993; Baldigo and Murdoch 1997) or there are nearby acid inputs (Brakke et al. 1988). It is not useful to specify a single value as the ambient water quality criterion for total or dissolved organic carbon. A more useful approach is to determine whether a particular anthropogenic activity (e.g., road construction, harvesting of forests) is causing a change in total organic carbon levels compared to historical conditions or to background conditions in nearby aquatic systems. We selected a 20% change in total and dissolved organic carbon for the criteria to ensure minimal impacts on aquatic biota. Smaller changes are not likely to be detected as significantly different from the stochastic variation often observed in British Columbia lakes and rivers (Figure 1).
The sampling design used to determine whether total or dissolved organic carbon has changed by about 20% as a result of an anthropogenic activity needs to be flexible. Some considerations include availability of historical data and/or reference sites, nature of the anthropogenic activity (e.g., point versus diffuse sources), natural seasonal variability, and various legal and economic issues. Short-term changes (e.g., <24 hours) can arise due to natural events (e.g., storms) and are unlikely to have serious impacts on aquatic communities. Therefore, when testing whether a particular anthropogenic activity has caused a failure to meet the organic carbon water quality guidelines, several samples (minimum n=5, preferably greater than 10) should be taken over a 30-day period and the 50th percentile calculated. The calculated 50th percentile for the site of interest should be compared to background data collected at similar times of the year and/or locations. A substantial
quantity of information has been collected on the levels of
TOC and DOC in British Columbia. This information is useful
for establishing background levels, although in some cases
the necessary data will not be available or will not be sufficient
to accurately determine reference conditions in the stream
system under investigation. In both of these cases, it will
be necessary to establish baseline conditions prior to the
implementation of developmental activities or establish appropriate
reference sites in upstream areas or nearby systems. It is
recommended that several years of background data from the
basin or site where management will occur and a similar set
of data from comparable, unmanaged sites be obtained.
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