Habitats of Freshwater Algae

Author's personal copy

Chapter 2

Habitats of Freshwater Algae John D. Wehr1 and Robert G. Sheath2 Louis Calder Center—Biological Station, Fordham University, Armonk, New York, USA. 2Department of Biological Sciences, California State University San Marcos, San Marcos, California, USA 1

Chapter Contents I. What are Freshwater Habitats? II. Lentic Habitats A. Major Lakes of North America B. Lakes and Ponds C. Small-Scale Lentic Habitats and Microhabitats D. Mutualistic, Commensal, and Symbiotic Habitats in Lakes E. Planktonic Habitats and Assemblages in Lakes F. Benthic Habitats and Assemblages in Lakes G. Wastewater Systems III. Lotic Habitats A. Brief Overview of the Geomorphology of Rivers B. Headwaters and Stony Streams C. Major Rivers of North America D. The River Continuum and Other Models E. Benthic Habitats and Assemblages in Rivers F. Planktonic Habitats and Assemblages in Rivers

13 14 14 17 20 21 23 25 29 30 30 30 33 35 37 40

42 IV. Wetland Habitats A. Functional Importance of Algae in Wetlands 42 B. Algal Diversity and Production in Freshwater Wetlands 43 V. Spring Habitats 45 A. Thermal Springs 45 B. Acid Springs 46 C. Karst Springs 47 VI. Subaerial Habitats 48 A. Soils 48 B. Epilithic and Endolithic Habitats 49 C. Plants and Animals 50 D. Snow and Ice 52 Acknowledgments 53 Literature Cited 53

I WHAT ARE FRESHWATER HABITATS? Algae occur in nearly all surface water bodies in every biome across the globe. Many of these environments are freshwater ecosystems. So what is fresh water? The distinction between freshwater and marine systems is not clear-cut. Oceans are clearly saline (~35 g of salts L−1) and most lakes relatively dilute (95% of the sewage-derived P (Hoffmann, 1998), and systems have been used to remove N and P from swine manure effluent (Kebede-Westhead et al., 2006). Algal turf-scrubbers use mass culture of algae to recover nutrients from dairy, swine, and other nutrient sources and use the harvested algae in slow-release green fertilizer, fermentation energy, and omega-3 products (Wilkie and Mulbry, 2002; Mulbry et al., 2008; Adey et al., 2013).

III LOTIC HABITATS Lotic habitats vary from small headwaters to huge river systems draining enormous watersheds. Running waters differ from lentic systems, characterized by directional flow, which joins a network of tributaries to form progressively larger rivers Hynes (1970). Streams and rivers are intimately connected with the surrounding landscape, which in turn regulates geochemical processes in the aquatic environment. Geomorphological features of river systems are described by Poole (2010) and Charlton (2013). Wetzel (2001) and Allan and Castillo (2007) describe the ecological properties of streams and rivers. An understanding of the habitats of algae in rivers begins with an understanding of the physical properties of the river and its watershed.

A Brief Overview of the Geomorphology of Rivers Streams and rivers form a network of connected tributaries whose hydrological features vary in predictable ways. Many of these features, such as discharge, substratum size, stream width, and depth, affect the species composition and productivity of lotic algae and their consumers. Larger substrata—cobbles and boulders—typify rapidly flowing streams, while sand and silt occur in slowly flowing rivers. The physical features of rivers are described in a system of stream orders, which assigns increasing numbers to streams when two tributaries of equal order join. The most widely adopted system (Strahler, 1957) defines a headwater stream or spring source with no permanent tributaries as first order, and the junction of two such streams a second-order tributary (Fig. 7). A second-order stream increases only when joined by another second-order stream, and so on. Larger-order streams are wider and longer segments, drain larger areas, and have a more gradual slope. The network of these stream segments forms a tree-like structure that is used in hydrological models to predict discharge, flood events, and quantity of suspended matter. While many features of drainage basins follow an ordered structure within a watershed, physical and ecological properties of streams vary greatly among locations. Current velocity, stream width, and substrata in a second-order stream in Nova Scotia differ that from a similar order, high desert stream in the western Great Basin, or a blackwater stream in Georgia (Minshall, 1978; Minshall et al., 1983; Benke et al., 1988; Meyer, 1990). Landscape factors such as climate, geology, riparian vegetation, land use, and nutrient sources exert major effects on biological assemblages. River channels also differ in the amount of interaction with their watershed, which is influenced by geological properties and degree of connection with groundwater (Dahm et al., 1998). Rivers may be geologically constricted, such as sections of the Hudson, Ohio, and Columbia, or have substantial floodplain interaction, such as the Upper Mississippi, Illinois, and Mackenzie. The floodplain includes wetlands, side channels, and floodplain lakes that are seasonally inundated during high flow. In areas of low relief, meandering rivers and those with more islands have greater littoral and floodplain interaction and complex current regimes (Fig. 8A). This increased habitat complexity offers habitat and refuges for larger, slowly growing algae. Small islands also form in lowland rivers as sandbars (Fig. 8B) or large stands of submersed angiosperms (Butcher, 1933; Holmes and Whitton, 1977). Within a given reach, there are alternating regions of erosion and deposition. In streams draining steeper slopes, differences in current velocity and substratum create regular, alternating patterns of riffles (rapid sections with larger substrata) and pools (deeper and slower sections), which are spaced more closely or widely depending on the slope of the stream gradient (Fig. 8C and D). Riffles have greater turbulence and concentrations of dissolved gases; pools form downstream of riffles, where organisms experience reduced current velocity. Sand and silt settle in pools under base flow, but during flood periods, these habitats become susceptible to scouring and erosion.

B Headwaters and Stony Streams A stream's origin is the headwater, which can be a spring, snowmelt, or small pool that drains downslope. Headwater streams are small and narrow and may be permanent or intermittent, depending on the water source (Meyer et al., 2007). Due to their small size, they are frequently are shaded by riparian vegetation (Fig. 9A), although in high alpine areas (Fig. 9B), as well as tundra, desert, grasslands, and cultivated landscapes, they can be open. Aquatic bryophytes often dominate shaded headwaters (Stream Bryophyte Group, 1999; Tessler et al., 2014). However, diatoms such as


Habitats of Freshwater Algae Chapter | 2 31

FIGURE 7 Structure of stream networks and tributaries within a watershed, indicating the numbering of stream orders.

FIGURE 8 Examples of physical structure and habitats in streams and rivers. (A) Meanders in Turkey River, a sinuous, lowland river, with alternating erosional and depositional habitats along the river's course (IA); (B) sandbars in the Wisconsin River, (WI); (C) cascading riffle-pool sequences in Coxing Kill, a high-gradient stream (NY); and (D) widely spaced riffle and long pools in a broad stream, Catskill Creek (NY). (Photos A and B by Louis J. Maher, Jr., with permission and photos C and D by J.D. Wehr).


32 Freshwater Algae of North America

FIGURE 9 Examples of headwaters and other stony streams. (A) A spring-fed headwater stream in a rhododendron-hemlock (Tsuga) forest (GA); (B) an unshaded alpine headwater stream (WA) draining a melting snow field (WA); (C) Tiorati Brook, a second-order mixed conifer-deciduous forest stream (NY); (D) Titicus Creek, a third-order stony stream in a deciduous forest during autumn leaf-fall (NY); (E) Big Brook, a humic-stained third-order boreal stream partially shaded year-round by black spruce (Picea mariana) (Nfld); (F) Quebrada Prieta, a second-order tropical forest stream (PR); (G) Green River, a fourth-order stream draining sagebrush habitat (WY); and (H) the Beaverkill, a broad third-order trout stream in mixed conifer-deciduous forest (NY). (Photo F by Todd Crowl, with permission and all other photos by J.D. Wehr).

Eunotia pectinalis, Meridion circulare, and Gomphonema parvulum were abundant in one shaded Appalachian headwater (Greenwood and Rosemond, 2005). The red alga Sheathia americana (as Batrachospermum boryanum) was abundant and persisted most of the year in a headwater woodland stream in Rhode Island (Hambrook and Sheath, 1991). A multi-biome experiment indicated that benthic algal production in most headwater streams was light limited; only streams in regions with open canopies were limited by nutrients (Johnson et al., 2009). Even at low densities, algal assemblages can be important in the diet of macroinvertebrates because of their high turnover rates (Mayer and Likens, 1987; McIntire et al., 1996; McNeely et al., 2007).



Algal assemblages change along the stream gradient (Rott and Pfister, 1988; Vavilova and Lewis, 1999). Second- to fourth-order streams in temperate deciduous forests can also be shaded, at least during late spring through summer, but after leaf-fall receive abundant sunlight for several months through winter (Fig. 9C and D). In drier climates or cultivated land, light limitation is less severe. Boreal streams (e.g., Fig. 9E) lack the seasonal canopy change but vary greatly in terms of pH and nutrient concentrations and may receive high amounts of dissolved organic carbon from the decay of coniferous vegetation (Buffam et al., 2007). Diatom assemblages differ among boreal streams according to difference in conductivity, pH, and nutrient status (Soininen et al., 2004). High humic, low pH, and nutrient-poor streams are typified by species such as Frustulia rhomboides, Tabellaria flocculosa, and Eunotia species, while neutral and more nutrient-rich streams include species such as Diatoma tenuis, Navicula trivialis, and Nitzschia palea. Macroalgal assemblages in boreal streams are among the most diverse in North America (Sheath et al., 1989). Climate warming is predicted to cause reduced water flow and decrease the ratio of base cations to strong acid ions in low-order boreal streams, furthering their acidification (Schindler et al., 1996). Tropical streams have received less attention, although several have been studied intensely for years. Quebrada Prieta (Fig. 9F), a second-order stream in Puerto Rico, receives heavy subsidies of terrestrial litter. Such systems are greatly influenced by shifts between rainy and dry seasons, hurricanes, changing land use, and even geothermal inputs (Pringle and Hamazaki, 1997; Ramírez et al., 2006). Organic matter is processed by freshwater shrimps, which along with macroinvertebrate insects and herbivorous fish also consume benthic algae (Pringle and Hamazaki, 1998; Crowl et al., 2001; Julius et al., 2005; Cross et al., 2008). Assemblages include species reported from temperate streams, plus taxa occurring only in tropical biomes, such as diatoms Hydrosera whampoensis and Planothidium salvadorianum, the green alga Cloniophora spicata, red algae Bostrychia moritziana and Caloglossa lereurii, and cyanobacteria Blennothrix ganeshii and Scytonematopsis contorta (Sheath et al., 1993a,b; Michels, 1998; Jiménez et al., 2005; Sherwood, 2006; Vaccarino and Johansen, 2011; Stephens et al., 2012). Broad stony streams in temperate biomes (Fig. 9G and H) can have extensive algal and angiosperm production, but can be affected by flooding following rain and snowmelt events. Filamentous forms such as Cladophora and Ulothrix are periodically scoured, while encrusting taxa such as the cyanobacterium Chamaesiphon and the phaeophyte Heribaudiella can persist year-round (Wehr and Perrone, 2003; Wellnitz and Rader, 2003; Fuller et al., 2008; Power et al., 2008).

C Major Rivers of North America Much larger rivers have different habitats and different factors influencing algal assemblages. As river size increases, planktonic forms predominate, although benthic algae can be abundant. The Mississippi-Missouri River system is the longest in the world (6970 km), with the third largest drainage area (3.3 × 106 km2), and is sixth largest in terms of discharge (18,390 m3 s−1; Milliman and Meade, 1983). The Ohio River, which joins the Mississippi at Cairo, Illinois (Fig. 10A), adds an additional 1580 km of river to this enormous system. Most major rivers have been substantially altered with navigation or hydroelectric dams, channelization, wetland removal, and pollutants (Sparks, 1995; Wehr and Descy, 1998). The Missouri River has six major impoundments over 1230 km (33%) of its length, while another 1200 km (32%) have been channelized. Only 35% (1330 km) of all river sections are estimated to be free flowing, although discharge is still influenced by reservoir conditions upriver (Hesse et al., 1989). No reservoirs are located on the Ohio or Mississippi Rivers, but low-head navigation dams and channelization throughout their lengths facilitate ship traffic. These structures create continuously flowing pools with reduced turbidity and longer retention times. Nonetheless, much of the Upper Mississippi has a connected floodplain, including wetlands, and backwaters (Fig. 10B), which facilitates the exchange of nutrients, organisms, and other materials within the river (Johnson and Hagerty, 2008). Each of these backwater habitats has particular geomorphological and chemical properties (Heiler et al., 1995; Strauss et al., 2011), which create diverse ecological conditions for aquatic vegetation and algal assemblages (Houser and Richardson, 2010; Kreiling et al., 2010). Restored backwater habitats in the lower Missouri have greater species richness and biovolumes of phytoplankton than the main channel (Dzialowski et al., 2013). The Illinois River (Fig. 10C) is a major tributary to the Mississippi (~440 km; 72,700 km2 drainage) and a prime example of a floodplain river with numerous side channels and backwater lakes that periodically connect and disconnect with the main channel, following the hydrograph. This turbid, low-gradient system has perhaps the longest history of scientific studies on plankton and other organisms on the continent, most notably the pioneering work by Kofoid (1903, 1908). Recent studies suggest that plankton assemblages vary strongly in space and time, strongly influenced by hydrological events (Wahl et al., 2008). The Mackenzie River (Fig. 10D) is another vast river system at 4240 km length, draining boreal, subarctic and arctic landscapes, countless wetlands and ponds, and several major lakes, including Great Bear Lake (Emmerton et al., 2007). Portions of the river remain frozen for many months of the year and receive peak runoff periods from ice melt in summer



FIGURE 10 Examples of large rivers in North America. (A) Landsat photo of junction of Ohio and Mississippi Rivers near Cairo IL; (B) Upper Mississippi River showing numerous backwater lakes and wetlands, near La Crosse (WI); (C) a section of the Illinois River with multiple side channels and flooded wetland forests contiguous with the main stem; (D) Mackenzie River near Inuvik, NWT (Canada); (E) middle section of the Columbia River in a constricted-channel section at Vantage Bridge (WA); (F) St. Lawrence River near Cornwall (ON), a section with rocky substrata and multiple small islands; (G) the Kansas River, a prairie river with sandbars and turbid water; and (H) Kissimmee River (FL), in a section in which restoration of the original floodplain has occurred. (Photo A courtesy of NASA; photos B, E, F, and H by J.D. Wehr; photo C by Andrew F. Casper, with permission; photo D by Warwick Vincent, with permission; and photo G by James H. Thorp, with permission).

(Vincent et al., 2008). These conditions contract the period during which algal production can develop, although recent data suggest significant warming in the Mackenzie basin (Vallières et al., 2008). The Columbia River (Fig. 10E), which flows ~2000 km from British Columbia through Washington and Oregon through geologically constricted, mountainous terrain, has been highly modified by hydroelectric dams, creating a series of slackwater pools and reservoirs, alternating with rapid regions. While dams are thought to reduce suspended particulate organic matter, particulate carbon during spring and summer is primarily composed of planktonic diatoms (Sullivan et al., 2001). It was suggested that abiogenic particles (turbidity) settle out in slackwater habitats and, when combined with longer



r etention times, create greater water clarity for phytoplankton production. This pattern was also observed upstream of navigation dams in the Ohio (Wehr and Thorp, 1997). The St. Lawrence River (Fig. 10F) receives water from the Laurentian Great Lakes and thus is strongly influenced by the geochemistry, organisms, and human activities in those basins. Algal biomass declines sharply as water empties from Lake Ontario into the river, due to increased turbulence and mixing (Basu et al., 2000). However, the freshwater section of the St. Lawrence includes several braided sections, with numerous fluvial lakes with extensive aquatic macrophyte beds that function to enhance both phytoplankton and zooplankton production in the river (Basu et al., 2000). Zooplankton and dreissenid mussels regulate phytoplankton densities in slackwater habitats (Thorp and Casper, 2003; Twiss and Smith, 2012). Most major rivers have substantial current speed in some sections and are well mixed, such that resuspension of sediments combined with mixing of cells through the water column can limit algal production (Thorp et al., 1994; Bukaveckas et al., 2011). This condition selects for algal species with rapid growths rates and efficient light harvesting capabilities (Reynolds, 1996). The Kansas River (Fig. 10G) is a prairie river, and although much smaller than major rivers (238 km), like many large rivers it periodically experiences high levels of turbidity. However, its shallow depth permits sunlight to reach the river bottom and sandbars serve to reduce turbidity, which supports greater plankton production (Thorp and Mantovani, 2005). The Kissimmee River (FL) was for centuries a free flowing, low-gradient, highly meandering river (166 km, >7800 km2 watershed), but was channelized and impounded in 1960s to support agriculture (Toth, 1993). The ecological damage that was created (e.g., loss of habitats, recued biodiversity, low dissolved oxygen) was motivation for the current restoration project (Fig. 10H), which aims to restore the habitats and ecosystem structure that existed before the river was channelized. Measurements of river metabolism in sections within reaches having restored flow conditions suggest metabolism rates similar to those in other free-flowing blackwater rivers (Colangelo, 2007).

D The River Continuum and Other Models For decades ecologists lacked broad conceptual models for describing and testing patterns in the structure and function of river ecosystems. Early ecologists described zones along a river with different physical conditions and habitats (Hawkes, 1975). But a river does not fall into discrete zones and instead varies continuously along its length. Rivers are open systems that transport materials and energy from one part of the watershed to downstream reaches. The combination of geomorphological conditions with corresponding changes in biological and chemical processes along a river gradient led to the development of the River Continuum Concept or RCC (Vannote et al., 1980; Fig. 11). The RCC considers lotic habitats and ecosystems as a network of streams with a continuum of longitudinally linked geomorphological and resource gradients. Biological communities respond to these longitudinal changes in many ways. Organisms, by their own activities further influence processes and assemblages downstream. This is evidenced by changes in the types of invertebrate consumers and algal taxa that occur in rivers of different sizes. The model was largely based on data from smaller, temperate forest streams and was broadened into a theory for rivers as a whole. Many other ideas and studies concerning nutrient spiraling, benthic invertebrates, fish production, and the influence of dams have emerged from the general framework laid out in the RCC (Newbold et al., 1981; Minshall et al., 1983; Ward and Stanford, 1983; Thorp and Delong, 1994). The importance of algae and other primary producers was also considered in the RCC. The metabolism of first- to third-order streams was viewed as largely dependent on external or allochthonous sources of terrestrial carbon for their metabolism. Hence, consumers in smaller streams were mainly shredders and collectors of coarse particulate matter. Algae were viewed as a minor component of food webs in low-order streams due to light limitation from heavy riparian shading and subsidies of terrestrial organic matter. The overall effect is a net heterotrophic system, in which ecosystem respiration exceeds in situ primary production (P:R 1.0 during spring and summer (Admiraal et al., 1994; Descy and Gosselain, 1994; Thorp and Delong, 2002).

E Benthic Habitats and Assemblages in Rivers Much research has been conducted on the composition, adaptations, physiology, and ecosystem importance of benthic algae in streams and rivers (Lock et al., 1984; Biggs, 1996; Steinman, 1996; Biggs et al., 1998; Stevenson et al., 2006, 2012). The focus in recent years has shifted from mainly descriptive studies to structural and functional analyses and their importance in food webs. There has also been wide use of benthic algal assemblages in bioassessment of river water quality (Hill et al., 2000; Walker and Pan, 2006; Jüttner et al., 2012; Kelly and Ector, 2012).

1 Benthic Algal Diversity, Composition, and Biogeography In streams, diatoms often comprise the dominant algal group in terms of species number and biomass (Blum, 1956; Douglas, 1958; Round, 1981; Kawecka, 1981; Biggs, 1996). Their diversity in species and growth form, including cells in stalked rosettes, tubes, filaments, upright frustules, prostrate cells, and mucilaginous masses enables them to colonize a variety of microhabitats. In more slowly flowing or less flood-prone systems, filamentous species like Melosira varians and upright, stalked forms such as Gomphoneis herculeana may predominate, along with filamentous nondiatom species (Biggs, 1996; Riseng et al., 2004). In very rapid water, firmly attached diatoms, such as Cocconeis placentula, Achnanthidium minutissimum, and Hannaea arcus, may occur with encrusting nondiatoms, such as Hildenbrandia rivularis, Gongrosira spp. and Chamaesiphon spp., and corticated forms like Lemanea spp. (Fritsch, 1929; Kawecka, 1980; Kann, 1978; Ledger et al., 2008). One of the fascinations of studying benthic algae in rivers is that many species (although not the majority) are macroscopic and recognizable in the field (Holmes and Whitton, 1977; Kann, 1978; Entwisle, 1989; Sheath and Cole, 1992; Sheath and Müller, 1997; Fig. 12). These include cyanobacteria, green and red algae, as well as chrysophytes, xanthophytes, and brown algae. Morphologies are diverse and include filamentous (Fig. 12A), blade-like (Fig. 12B), mucilaginous (Fig. 12C and D), cartilaginous (Fig. 12E), erect tufts (Fig. 12F), tubular, encrusting (Fig. 12G), turf-like, and plant-like (Fig. 12H) thalli. In some instances, diatoms, including Eunotia, Fragilaria, and Melosira, and stalked forms, such as Didymosphenia geminata and Cymbella spp., produce large masses and may be recognized by their gross appearance, but they are usually identified using microscopy (Holmes and Whitton, 1981; Steinman and Sheath, 1984; Sheath and Cole, 1992; Whitton et al., 2009). Some macroalgae have limited distributions or are at least infrequently reported. Prasiola mexicana, a seaweed-like green alga (Fig. 12B), thus far has been recorded in streams located only in the Arctic tundra and western coniferous biomes (Sheath and Cole, 1992), and the gelatinous cyanobacterium Nostochopsis lobata (Fig. 12D), has been recorded only occasionally in both North America and Europe, despite its macroscopic size (Smith, 1950; Sherwood, 2006; Moreno et al., 2012). Thorea violacea is a large red alga (reaching >1 m in length) thought to be restricted to streams in warmer biomes or in temperate areas only during summer (Smith, 1950; Sheath and Hambrook, 1990). However, it was discovered growing profusely in the upper Hudson River in cool (15 °C), rapidly flowing water (Pueschel et al., 1995). Few long-term studies of benthic stream macroalgae exist that may help to explain the dispersal patterns of these organisms. In one example of a 40+-year database, Holmes and Whitton (1977, 1981) characterized species as previously overlooked, currently increasing, or recently extirpated from the River Tees (U.K.). No such studies are known from North America. Ecological endemism in some species of river algae may be caused by the spread of marine taxa into estuarine environments, followed by adaptation to lower salinity. Such may be the case with genera such as Audouinella, Hildenbrandia, Prasiola, and Ulva (formerly Enteromorpha): genera that occur in rivers and have both marine and freshwater species (Flint, 1955; Sheath and Cole, 1980; Sheath et al., 1993a,b; Hamilton and Edlund, 1994). However, other species appear to be human-accelerated invaders from other freshwater systems. The red alga Bangia, which was first recorded in



FIGURE 12 Examples of different forms of benthic macroalgae in streams and rivers. (A) Zygnema sp., simple, flexible filaments (arrow = direction of current); (B) Prasiola mexicana, flat, flexible blade or thallus; (C) Nostoc verrucosum, firm mucilaginous colonies; (D) Nostochopsis lobata, soft gelatinous colonies; (E) Lemanea fluviatilis, corticated tubes; (F) Audouinella hermannii, upright tufts; (G) Heribaudiella fluviatilis, tightly adhering crusts; and (H) Chara zeylanica, plant-like morphology (scale = 5 cm for A–E and G; 1 cm for F; 50 cm for H). (Photos A, C, and H by J.D. Wehr; photos B and E by R.G. Sheath; and photos D, F, and G by Kam Truhn, with permission).

North American fresh waters in the 1960s (Sheath et al., 1985), is not closely related to any marine populations (based on rbc-L, RuBisCo spacer, and 18s rDNA sequences) and has strong affinities with freshwater European populations (Müller et al., 1998). These data suggest a vector, such as a ship's ballast water, enabled invasion, rather than a gradual spread and adaptation. Biogeographic data indicate that green algae are the most common group of stream macroalgae across all biomes in North America. Based on 1000 stream segments studied, lowest species diversity was observed in streams in arctic tundra and greatest in boreal forests (Sheath and Cole, 1992). There was no increase in macroalgal species diversity from the arctic to the tropics, contrary to marine taxa. Arctic streams tend to have more species of macroalgal cyanobacteria, while tropical streams have more species of Rhodophyta (Sheath and Cole, 1992; Sheath et al., 1996; Sheath and Müller, 1997).



2 Factors Regulating Benthic Algae in Rivers Several reviews and studies (Blum, 1956; Whitton, 1975; Biggs, 1996; Stevenson, 1997; Potapova and Charles, 2002) discuss many ecological factors and their effects on benthic algae in rivers. These multiple factors interact to affect algal growth and survival, thus multivariate analyses have been useful in determining the key environmental variables affecting species composition, succession, and distributions (e.g., Wehr, 1981; Leland and Porter, 2000; Griffith et al., 2002; Cardinale et al., 2006). However, interpretations based on these analyses can be affected by sampling effort (Cao et al., 2002), and experimental studies are needed to identify the mechanisms responsible for patterns that have been observed (Nelson et al., 2013). Some generalizations, however, can be made. In all stream and river systems, there are proximate variables that affect organisms, such as light availability or nutrient supply, and larger-scale factors such as climate, watershed features, and land use practices, which drive proximate variables (Stevenson, 1997). Biggs (1996) has proposed a disturbance-resource-supply-grazer model that categorizes the controls on benthic algal production in terms of two processes, factors affecting (1) biomass accrual and (2) biomass loss. An energetic balance sheet is constructed for each side of the ledger that can be used to predict to algal production and species composition. Biomass accrual increases as a function of resource supply, while biomass losses are a function of disturbance and grazing. In streams with infrequent, low-intensity floods and modest grazing intensity, the model predicts biomass accrual dominates to a level dictated by resources. Under low resource supplies, growth continues at lower rates, favoring adnate and turf-like assemblages dominated by filamentous cyanobacteria (e.g., Schizothrix, Phormidium, Tolypothrix), red algae (e.g., Audouinella), and benthic diatoms (e.g., Epithemia, Navicula). Under similar hydraulic and grazing conditions but greater nutrient supply, greater biomass of filamentous taxa, such as Cladophora, Ulothrix, or Melosira, is expected. For each combination of + and − factors, specific predictions can be made about which variables regulate algal production and taxonomic composition. A large study offers strong evidence for both N and P control on benthic algal biomass across many temperate rivers and streams in North America (Dodds et al., 2002). A section of the Colorado River below Glen Canyon Dam is a system in which losses due to disturbance are critical. In this system, nutrients and sunlight are generally nonlimiting, but release of water from the dam (greater disturbance) decreased the biomass and relative importance of Cladophora glomerata and reduced total benthic primary production (Blinn et al., 1998). The physiognomy of epiphytic algae changed from an upright assemblage (Diatoma vulgare, Rhoicosphenia abbreviata) to closely adherent forms (Achnanthidium spp., Cocconeis pediculus) (Hardwick et al., 1992). Losses due to herbivorous invertebrates have also been widely shown, but losses due to grazer activity come from physical disturbance by invertebrates as well as consumption (Allan and Castillo, 2007). Further, ingestion and assimilation of benthic algae vary widely (30-70%), depending on the species of both the alga and consumer (Lamberti et al., 1989; Pandian and Marian, 1986). Invertebrates may indirectly affect algal physiognomy as well. The morphology of Nostoc parmelioides colonies in streams is altered from spherical to ear-shaped colonies by the presence of an endosymbiotic midge larva (Ward et al., 1985), and only ear-shaped colonies exhibit greater photosynthesis and N2-fixation rates with greater current velocity (Dodds, 1989). Changes in biomass, composition, and physiognomic stricture have effects on benthic food webs, because many invertebrates find an adherent community less easily grazed (Colletti et al., 1987). The Rivière de L'Achigan (Quebec) is an unshaded stream interspersed by a chain of small lakes that alter flow conditions immediately downstream (lower disturbance) (Cattaneo, 1996). Biomass and species composition of benthic algae on gravel varied inversely with distance from lake outlets, yet this impoundment effect was unimportant for algae colonizing boulders. Only boulders supported assemblages of large filamentous and plumose forms such as Draparnaldia (Chlorophyceae), Stigonema (Cyanobacteria), and Batrachospermum (Rhodophyta). In Big Sulfur Creek (CA), algal biomass in low-grazer conditions declined by 75% when under greater canopy cover, but was unaffected by light availability at normal grazer densities (Feminella et al., 1989). The food web—resource supply interaction in Big Sulfur Creek is complicated by the fact that densities of trichopteran grazers declined significantly in shaded conditions, while other species were unaffected. Reductions in light availability to Kingsley Creek (NY) caused significant reductions in epilithic algal biomass and herbivorous mayflies but did not affect densities of filter-feeding blackfly larvae, which consumed detritus (Fuller et al., 1986). Differences in irradiance also affect the chemical composition of algal assemblages, including changes in protein, total lipids, and fatty acid composition (Steinman et al., 1988). Fatty acid content of algal assemblages is also affected by differences in the combination of light levels and aqueous nutrients (Cashman et al., 2013), which is important for benthic consumers (Torres-Ruiz et al., 2010). In general, cyanobacteria have different complements of fatty acids, with greater oleic, linoleic, and linolenic acids than diatoms (McIntire et al., 1969; Torres-Ruiz et al., 2007). Current velocity is of great importance in stream habitats. Algal species that colonize microhabitats with rapid current are firmly attached to substrata using rhizoidal or holdfast-like structures (Israelson, 1949; Lowe and Hunter, 1988). Greater current velocity provides a continuous replenishment of nutrients from upstream and a steeper diffusion gradient



near the cell surface (Whitford, 1960; Horner et al., 1990). Species restricted to habitats with strong current velocity, such as Hydrurus foetidus and Lemanea fluviatilis, may have a higher metabolic demand for nutrients, which must be met by a greater physical supply. But there is no simple relation between current velocity and algal metabolism or growth rate, because greater current speed can also decrease biomass through scouring (Borchardt et al., 1994; Stevenson, 1996). And in the case of Hydrurus, cold temperature is also a limiting (required) factor (Klaveness and Lindstrøm, 2011). A study of parallel streams draining the same reservoir observed that the regulated stream (less disturbance) had greater cover of diatoms and Prasiola fluviatilis, while the stream lacking flood control (greater disturbance) had greater cover of Hydrurus foetidus and Ulothrix zonata and greater species richness but lower biomass of diatoms (Kawecka, 1990). Current velocity affects the growth form of individual algae, such as in Cladophora, which can develop compact tufts with narrow branching in greater current velocity and widely branched, plumose forms in calmer flow (Whitton, 1975; Dodds and Gudder, 1992). Type of substratum is also important. More than five decades ago, Douglas (1958) recognized that different sizes of stones supported different densities and species of epilithic algae, which was likely the result of differences in their susceptibility to flood disturbance. A comparison of epilithic assemblages in streams with different substratum sizes and nutrient conditions demonstrated that most of the variation in algal biomass was explained by total-P and seston, but size of substratum also exerted a significant effect (Cattaneo et al., 1997). Stones that are disturbed or scoured during floods may still have algal crusts or propagules remaining on their surfaces (Power and Stewart, 1987). The degree to which algae can recolonize disturbed substrata is a function of their resilience (Steinman and McIntire, 1990), through immigration and greater growth rates, or resistance, as influenced by the species' morphology and community physiognomy (Peterson, 1996). Diatom immigration onto bare substrata will increase with either reduced current speed or greater surface complexity (Stevenson, 1983). Substrata conditioned with simulated mucilage (agar coating) were colonized twice as rapidly as clean surfaces, but responses were species-specific; some increased (Navicula gregaria, Synedra ulna), while others declined (Achnanthidium minutissimum, Diatoma vulgare) or were unaffected (Diatoma tenuis, Gomphonema olivaceum). A matrix of organic matter and bacteria probably facilitates colonization (Karlström, 1978; Korte and Blinn, 1983; Sheldon and Wellnitz, 1998). A comparison of benthic diatom assemblages colonizing natural rocks, sterilized rocks, and clay tiles in Fleming Creek (MI) recorded greater total densities and species diversity on natural rocks (Tuchman and Stevenson, 1980). Otherwise identical substrata constructed from sandstone had significantly greater diatom density than two other naturally occurring rock types (basalt, limestone) in Oak Creek (AZ), although species composition and diversity were similar (Blinn et al., 1980). Complexity, roughness, and orientation of substrata also matter, with greater abundance and diversity of diatoms on rock surfaces with deeper crevices (Bergey, 1999; Murdock and Dodds, 2007). In Mack Creek (OR), greater biovolume and diversity of benthic algae were observed colonizing pieces of wood than on clay tiles (Sabater et al., 1998). Certain taxa, such as Encyonema minutum, Hannaea arcus, and Zygnema sp., were more abundant on wood, while closely adherent forms, such as A. minutissimum and Planothidium lanceolatum, were more abundant on clay tiles.

F Planktonic Habitats and Assemblages in Rivers Although benthic algae typically dominate smaller streams and rivers, phytoplankton are important in large rivers (Reynolds and Descy, 1996). There is a long history of studies on river phytoplankton dating back to at least the 1890s, when Zacharias (1898) coined the term potamoplankton to refer to the suspended organisms in flowing waters. In North America, phytoplankton assemblages of rivers have been studied since the early years of limnological research, including the Illinois (Kofoid, 1903, 1908), Mississippi (Reinhard, 1931), Ohio (Eddy, 1934), San Joaquin (Allen, 1921), and Sacramento (Greenberg, 1964) rivers. Much of the early research focused on whether a true phytoplankton assemblage (populations that survive and reproduce within rivers) actually exists, as opposed to dislodged benthic forms or plankton washed in from lakes in the basin. Indeed, phytoplankton in most rivers consists of all three components in varying proportions (Reynolds, 1988). Benthic algae can become suspended when washed out from sediments, plants, or other substrata (Leland and Porter, 2000). A meta-analysis suggests that about 50% of suspended algal taxa in rivers are either benthic or resuspended (Rojo et al., 1994). Among diatoms, most raphe-bearing species are likely nonplanktonic, but it is difficult to distinguish true potamoplankton from those originating from lakes (Reynolds, 1988). One distinction may be in species' responses to flow regime and other physical factors, as described for Ohio River phytoplankton (Peterson and Stevenson, 1989; Wehr and Thorp, 1997). Abundances of many species were negatively related to discharge—an indication they were not benthic in origin. However, during summer low flow, significant populations of colonial cyanobacteria (e.g., Aphanocapsa saxicola) and unicellular diatoms (e.g., Stephanocyclus meneghiniana) developed downstream of these tributaries, suggesting that some populations may have originated from outside the main river. Phytoplankton production in main channel regions of rivers is highly influenced by river stage, particularly large rivers with connected floodplain lakes and other backwater habitats (Moss and Balls, 1989; García de Emiliani, 1997; Tockner et al., 2000). This dynamic is referred to as a “flow-pulse,”



in which inputs from backwater nutrients and local algal production contributes to high spring production in some rivers (Tockner et al., 2000; Ochs et al., 2013). The fact that chlorophyll-a concentrations in large rivers often vary inversely with discharge (Schmidt, 1994) suggests that potamoplankton may not be derived primarily from benthic habitats (Ruse and Love, 1997). Alternatively, phytoplankton biomass can become diluted during high-flow periods. In general, smaller forms are more successful members of the potamoplankton in the main channel of larger rivers, perhaps due to rapid growth rates (Reynolds, 1988; Rae and Vincent, 1998). Diatoms are clearly the most diverse and abundant group, with Cyclotella, Cyclostephanos, and smaller species of Stephanodiscus especially common in larger rivers (Reynolds and Descy, 1996; Leland et al., 2001; Descy et al., 2012). Algal flagellates rarely achieve large numbers in river plankton, except some cryptomonads, euglenophytes, chrysophytes, and members of the Volvocales in low-flow sections (Baker and Baker, 1981; Kiss and Kristiansen, 1994; Stevenson and White, 1995; Leland et al., 2001). Two major limitations to survival and growth of river phytoplankton are the continuous removal of organisms by downstream flow and mixing within the water column, placing cells in variable and often aphotic light fields. Hence, most studies conclude that riverine phytoplankton production is controlled by discharge (Baker and Baker, 1979; Soballe and Kimmel, 1987; Cole et al., 1992; Reynolds and Descy, 1996). Assuming no other limiting resources, rivers must be long enough and/or the flow rate low enough for net positive algal growth rates. This principle was demonstrated in early studies on the Sacramento River, in which peaks of abundance in potamoplankton became progressively more pronounced further downstream, a region that provided reduced current velocity and longer times for populations to develop (Greenberg, 1964). A similar increase was seen along the Rhine (Germany-Netherlands; de Ruyter van Steveninck et al., 1992). In the lowland River Spree (Germany; Köhler, 1993, 1995), phytoplankton biomass declined in midriver in response to increased turbidity and Fe-induced precipitation of P, but then increased further downstream, as a result of impoundments and flow regulation. A different longitudinal pattern is seen in the St. Lawrence River, in which a gradient of increased phosphorus and reduced current velocity is counterbalanced by greater suspended matter, causing a net decrease in river plankton densities downstream (Hudon et al., 1996). Reynolds (1988, 1995) has discussed in detail the contrary and complex influences of discharge and algal growth rates. Despite these limitations, accumulations of phytoplankton frequently develop in summer in many backwaters and some main channel habitats. With greater nutrient supplies, surface blooms of cyanobacteria (Microcystis, Anabaena, and Aphanizomenon) have become increasingly common (Baker and Baker, 1979; Krogmann et al., 1986; Paerl and Bowles, 1987). Nutrient limitation is less common in larger rivers than in lakes (Reynolds and Descy, 1996; Wehr and Descy, 1998), although exceptions exist (Descy et al., 2012). In conditions of nutrient surplus, the principal limiting factor is generally discharge, which regulates dilution rates, turbidity, and mixing of cells within the water column (Reynolds and Descy, 1996). However this is not the case in blackwater rivers, where DOC-derived light attenuation is more important than flow (Meyer, 1990). Algal production in larger rivers can be substantial despite frequent turbidity and continuous mixing of algal cells within the water column (Descy et al., 1987, 1994; Reynolds and Descy, 1996). There is a delicate balance between phytoplankton production and respiratory losses during periods of higher turbidity (Descy et al., 1994; Reynolds and Descy, 1996). In freshwater sections of the Hudson River, turbidity is further complicated by tidally driven mixing, resulting in a net heterotrophic balance for most of the year (Cole et al., 1992). Following invasion of the Hudson by zebra mussels, algal biomass declined from a mid-summer mean of about 30-5 μg chlorophyll-a L−1, while species composition shifted from colonial cyanobacteria to diatoms (Caraco et al., 1997; Smith et al., 1998). In the River Spree, a positive autotrophic balance has been recorded in spring (mainly diatoms), but not summer, when cyanobacteria were dominant (Köhler, 1995). How large river systems maintain large phytoplankton populations over the year is still something of a mystery, but main channels often receive subsides of algae and nutrients from tributaries, wetlands, or backwaters (Owens and Crumpton, 1995; Reynolds, 1996; Ochs et al., 2013). Species composition may provide a clue: Slower growth rates of larger colonial species have higher respiratory costs for maintaining populations than do smaller centric diatoms, but may be stable if ineffectively grazed by small-bodied zooplankton (Gosselain et al., 1998). Only zebra mussels, which utilize a wider particle size range, may crop these larger forms. In the upper Mississippi and other large rivers, higher levels of nutrients coupled with less turbid conditions enable high levels of phytoplankton production to be sustained for several months of the year (Baker and Baker, 1979; Descy et al., 1987; Admiraal et al., 1994; Lange and Rada, 1993). This enhanced primary production influences biogeochemical processes in rivers, including dissolved O2 (Köhler, 1995; Reynolds and Descy, 1996), Si (Admiraal et al., 1993), and DOC (Wehr et al., 1997). Phytoplankton cells are an important food source for zooplankton in rivers, even if grazers do not regulate algal biomass or production as effectively as in lakes. Grazing pressure is thought to be less important because the zooplankton community is usually dominated by small-bodied cladocerans and rotifers (Thorp et al., 1994; Köhler, 1995). Rivers select for small-bodied zooplankton because of their ability to grow rapidly enough to compensate for downstream losses (Viroux,



1997; Baranyi et al., 2002). Biomass and density of zooplankton in larger rivers may also be less than in lakes (Pace et al., 1992; Thorp et al., 1994). While discharge and turbidity typically drive phytoplankton production in rivers on an annual basis, zooplankton grazing may still be an important loss factor during summer low-flow periods and alter the size structure of the algal assemblages (Gosselain et al., 1994, 1998). Models designed to predict phytoplankton production in large rivers have mostly been devised for specific conditions, such as the influence of temperature and irradiance on phytoplankton production in one reach of the Great Whale River in northern Quebec (Rae and Vincent, 1998). Light and temperature explained between 74% and 98% of the variation in photosynthetic activity in this subarctic river. A large model developed for the Rhine, which included effects of irradiance, light attenuation, flow, nutrients, phytoplankton biomass, and grazing, was successful in predicting the fate of algal production, although some parameters (e.g., zooplankton grazing) were based on data from lakes (Admiraal et al., 1993). Efforts to develop nutrient-algal biomass models have been less effective than in lakes (cf. Mazumder and Havens, 1998), owing to weaker relationships between N or P and chlorophyll-a, and complicating effects of discharge and turbidity (Van Niewenhuyse and Jones, 1996). In general, models that have incorporated several physical and chemical variables along multiple river sites indicate that the main factors regulating phytoplankton dynamics are hydrological and meteorological (Billen et al., 1994; Descy et al., 2012). More difficult is predicting changes in species composition of phytoplankton communities in rivers. Algal metrics have been developed for periphyton (benthic algae) in major rivers in the United States and have proven effective in assessing nutrient sources from agricultural and urban landscapes (Porter et al., 2008). Data indicate that the abundance and taxa richness of resuspended benthic forms was greater in rivers with elevated turbidity and nutrient concentrations. Using these patterns to understand water quality is another important goal. A phytoplankton-based index was developed for lowland rivers in Germany, which was in part based on older saprobic and diversity measures (Wu et al., 2012). Newer methods are being developed in North America (e.g., Flotemersch et al., 2006; Royer et al., 2008; Maret et al., 2010), but wide testing is needed as elevated nutrients and taste and odor episodes become more frequent in major rivers.

IV WETLAND HABITATS Wetlands regulate nutrient fluxes between terrestrial and aquatic systems, serve as nurseries for fisheries, and are among the most productive and threatened ecosystems worldwide (Mitsch and Gosselink, 2007). There are oligotrophic and dystrophic wetlands, such as peat bogs, wet savannas, and shorelines in sand plains. Freshwater wetlands occur from arctic to tropical biomes. Many are situated in the upper littoral zone of lakes and rivers, and include marshes, fens, alpine meadows, cypress swamps, and forested lowlands. Their unifying features include saturated soils, fluctuating water levels, an accumulation of detritus and organic matter, and vegetation adapted to wet conditions (Goldsborough and Robinson, 1996; Mitsch and Gosselink, 2007).

A Functional Importance of Algae in Wetlands Most wetland algae are benthic, loosely associated with emergent plants (metaphyton), attached to plants (epiphyton), or colonize sediments (epipsammon/epipelon). Suspended forms are mostly dislodged from various surfaces. A five-year study of epipelic, epiphytic, metaphytic, and planktonic primary production in Delta Marsh, Manitoba, determined that metaphyton contributed roughly 70% of the total algal productivity, compared with only 6% for phytoplankton (Robinson et al., 1997). Values for algal productivity (400-1100 g C m−2 y−1) over the year are comparable to or exceed that of the emergent macrophytes present (aboveground: 100-1700 g C m−2 y−1). Among several constructed wetlands in Illinois, benthic algae are estimated to contribute between 1% and 65% of the total system primary production (Cronk and Mitsch, 1994). However, algal assemblages in low-nutrient and humic wetlands are much less productive (Murkin et al., 1991; Goldsborough and Robinson, 1996). Algal production is important in wetlands for invertebrate consumers that preferentially consume algal material over live or detrital macrophyte tissues (Campeau et al., 1994; Goldsborough and Robinson, 1996). Algal material may be especially important in winter when emergent macrophytes are dead or senescent (Meulemans and Heinis, 1983). In a wetland along western Lake Superior, δ13C data suggest that entrained algae are a key primary food source in the grazing food web (Keough et al., 1998). Using plant pigments, Bianchi et al. (1993) estimated that benthic diatoms comprised a major food source for invertebrates in Hudson River wetlands and, when combined with lower C:N ratios, may be a better food resource than detritus. Nutrient additions designed to enhance algal biomass in a Manitoba wetland resulted in greater densities of cladocerans and copepods in near-shore water, as well as benthic snails and chironomids (Gabor et al., 1994). Wetlands and associated algal assemblages in the littoral zone of lakes function in nutrient exchange via uptake, transformations, and N2-fixation (Mickle and Wetzel, 1978; Moeller et al., 1988; Scott et al., 2005). Near-shore wetlands of lakes may



also be locally important areas of system productivity (Cooper et al., 2013). During wetland plant decomposition, attached algae may enhance their breakdown (Neely, 1994), although DOC released from some macrophytes during this process can inhibit algal growth (Cooksey and Cooksey, 1978).

B Algal Diversity and Production in Freshwater Wetlands Algal assemblages in wetlands like those in lakes are influenced by nutrient conditions, but factors such as soil type, water level, plant composition, and degree of exchange with other water bodies are also important (Goldsborough and Robinson, 1996). Because many wetlands are major habitats for waterfowl, their feces can be a major nutrient source for benthic and planktonic algae (Purcell and Goldsborough, 1996). In the Everglades (Fig. 13A–D), epiphytic algae colonize macrophyte and sediment surfaces and form thick carbonate-rich accretions or “sweaters” on plant stems (Gaiser et al., 2011). Algae can mediate calcium phosphate (apatite) deposition in wetlands by scrubbing phosphorus from waters (net retention) as they flow-through wetlands (Dodds, 2003). Species composition and production vary with species of macrophyte, water level, nutrient inputs, and degree of CaCO3 incrustation (Browder et al., 1994; Gaiser et al., 2011). Filamentous cyanobacteria, including Scytonema, Schizothrix, Oscillatoria, and Microcoleus, are abundant in carbonate-rich habitats. Filamentous green algae (Spirogyra, Bulbochaete, and Oedogonium), desmids, and diatoms (Cymbella, Gomphonema, and Mastogloia) are common in slightly less calcareous conditions. Algae represent between 30% and 50% of primary producer biomass in these systems, and their activity is apparent in the large diurnal changes in dissolved O2 and CO2. Benthic algae help regulate wetland water quality, especially P loading from agricultural and urban runoff, which has increased in recent decades (McCormick and Stevenson, 1998; Pan et al., 2000). Calcareous marshes in Belize have a similar cyanobacterial flora, with greatest biodiversity in systems with moderately high conductivity (1000-2000 μS cm−1; Rejmánková et al., 2004). Littoral wetlands and marshes in Farmington Bay of Great Salt Lake (Fig. 13E) receive freshwater and nutrients from the Jordan River and experience elevated algal biomass, often >100 μg L−1 as chlorophyll-a (Marcarelli et al., 2006). These marshes are based on an organic particle + benthic algal-based food web that supports brine flies (Ephydra cinerea) and that is weakly linked with the phytoplankton-based web of the main lake (Belovsky et al., 2011). In more saline littoral habitats, benthic cyanobacteria (Aphanothece utahensis) create large reef-like stromatolites or biostromes that cover large areas and serve as habitat for brine flies (Carozzi, 1962; Wurtsbaugh, 2009). These structures accumulate selenium and mercury, which may be passed on through the food web (Wurtsbaugh et al., 2011). Vernal pools (Fig. 13F) are widespread in forested region across the continent, and their duration as aquatic habitats varies from weeks to months depending on climate and the surrounding water table (Brooks, 2004; Schneider and Frost, 1996). Provided sufficient light, in late spring the water surface can become covered with mats of filamentous and gelatinous green algae, taxa such as Mougeotia, Microspora, Oedogonium, and Ulothrix, which would eventually dry up later in the summer. One study has demonstrated that the type of leaf litter that falls into the pools can influence to some degree their algal assemblages (Verb et al., 2001). Pine litter was demonstrated to depress pH but increase NO3− concentrations, relative to deciduous litter. There is evidence that the metabolism of vernal forest ponds may be net heterotrophic and that leaf litter suppresses algal production (Rubbo et al., 2006). However, despite the importance of vernal pools as habitats for many amphibians, surprisingly little work has been conducted on the role of algal assemblages in food webs in these temporary ponds. Larger forested wetlands and wetland ponds in northern locations that are more open to full sunlight (e.g., Fig. 13G) tend to have a diverse assemblage of aquatic macrophytes present and exhibit a much greater representation of phytoplankton species (Meyer and Brook, 1968, 1969a,b; Bland and Brook, 1973; Ngo et al., 1987; Johnson et al., 2007). An important strategy for the success of algal species in wetland habitats is their ability to tolerate variations in water level and desiccation (Goldsborough and Robinson, 1996). Algae that occupy a variable moisture regime must have adaptations to tolerate extremes of conditions. Some epipelic desmids (Closterium and Micrasterias species) are capable of surviving extended periods of drying and darkness (Brook and Williamson, 1988). Other filamentous forms (e.g., Oscillatoria, Lyngbya, Oedogonium) may form thick mats during the open (flooded) state, which may protect algal cells during a later dry phase. Fluctuations in water level (hydroperiod) also influence nutrient exchange between water and sediment, with desiccation and subsequent inundation resulting in large nutrient pulses (Steinman et al., 2012). These fluctuations can also trigger increases in benthic algal abundance and biovolume, as well as shifts in species composition (Rober et al., 2013). Bogs are a special class of wetland habitat whose water retention and chemistry are influenced by Sphagnum. Most have lower pH (4.0-5.5), low Ca2+, are poor in nutrients, and have high levels of dissolved organic matter (humic acids), which casts a yellow or brown stain to the water (Fig. 13H; Gorham et al., 1985). Bogs are scattered throughout North America and are common in arctic, boreal, temperate, and coastal regions, with a few in desert and tropical areas (Brooks and Deevey, 1963; Yung et al., 1986; Schalles and Shure, 1989). Lindeman's pioneering study of Cedar Bog Lake, Minnesota (Lindeman, 1941, 1942) helped launch the trophic-dynamic concept in ecology and served as the basis



FIGURE 13 Examples of wetland habitats. (A) Landsat image of southern Florida showing the extensive connected streams, marshes and ponds of the EvergladesvLake Okeechobee wetland system; (B) wetland pond system in the Everglades; (C) floating periphyton mats from Everglades wetland; (D) close-up of calcareous periphyton “coats” on wetland plant stems (scale = 2 cm); (E) wetlands in Farmington Bay bordering Great Salt Lake (UT); (F) vernal pool in a northeastern U.S. hardwood forest (NY); (G) wetland ponds, Itasca State Park (MN); and (H) humic-stained boggy pool on Vancouver Island (BC). (Photo A courtesy of NASA; photos B–D by Don Chamberlain; and photos E–H by J.D. Wehr).

for energetic studies of freshwater food webs and the development of the ecosystem concept. Algal assemblages may be species poor, although diversity, especially those of the desmids, is much greater in systems connected to lakes or streams (Woelkerling, 1976; Hooper, 1981; Mataloni and Tell, 1996). A tangle of filamentous green algae and desmids is common, although desmids, while diverse, are rarely important in terms of algal biomass. Data from 31 eastern bog systems demonstrated that algal species richness increases (especially desmids) with proximity to the Atlantic coast and is less in systems with greater color and lower pH (Yung et al., 1986). The flagellate Gonyostomum (Raphidophyceae) and the red alga Batrachospermum turfosum are two species characteristic of boggy wetlands (Yung et al., 1986; Sheath et al., 1994). Diatoms are generally less diverse than desmids, but certain species are especially common, including Brachysira brebissonii, Frustulia saxonica, Eunotia exigua, Kobayasiella subtilissima, and several Stauroneis spp. (Bruno and Lowe, 1980; Kingston, 1982; Cochran-Stafira and Andersen, 1984; Mataloni and Tell, 1996). Cyanobacteria have low diversity



in these systems, but species such as Aphanothece microscopica, Stigonema mamillosum, and species of Chroococcus, Geitlerinema, Hapalosiphon, Merismopedia, and Nostoc are commonly reported (Hooper, 1981; Dawes and Jewett-Smith, 1985; Yung et al., 1986; Mitchell et al., 2003; Jassey et al., 2011).

V SPRING HABITATS Springs have been the focus of study by phycologists and microbiologists for more than a century, drawn to these extreme and unusual environments as living laboratories for physiological, ecological, and evolutionary studies (Round, 1981; Papke et al., 2003; Heffernan et al., 2010). Springs by definition are isolated habitats, and as such are analogous to islands (Glazier, 2009), although some are connected to larger water bodies, which may serve as a means for propagule exchange. Recent advances in DNA methods and metagenomics, combined with an interest in characterizing biodiversity and microbial endemism, has stimulated new research (Souza et al., 2008; Breitbart et al., 2009).

A Thermal Springs Thermal springs and streams (hot springs) are extreme environments located in geologically active regions, and whose temperatures range from >35 to 110 °C. Among the best known are the thermal springs, geysers, and fumaroles in Yellowstone National Park. In some, travertine can form, with aragonite crystals forming “shrubs” with bacterial assemblages near the spring source (Pentecost, 1990) and cyanobacteria in slightly cooler waters some distance from the source (Fig. 14A). Some thermal springs act as tributaries and exert their chemical and thermal effects on lakes and rivers, as in Yellowstone Lake and streams and lakes in Costa Rica (Pringle et al., 1993; Theriot et al., 1997). The geology, chemistry, and organisms colonizing hot springs have been reviewed in detail by Castenholz and Wickstrom (1975), Brock (1986), and Ward et al. (2012). Temperatures can be fairly constant near the source, but can range from about 110 °C to just above ambient, depending on the temperature and volume of thermal water, distance from the source, and volumes of nonthermal surface water. Temperature is not the only extreme condition for thermal organisms; most have elevated concentrations (50-150 mg L−1) of inorganic ions (Ca2+, Mg2+, Na+, HCO3−, SO42−, Cl−, Si, and H2S) and elevated (8-10) or acidic (2-4) pH. These conditions select for highly adapted organisms, especially chemoautotrophic and heterotrophic bacteria in very hot (>70 to ~94 °C) conditions (Brock, 1985). Among photosynthetic organisms, cyanobacteria are by far the most common, with an upper limit of ~74 °C (Ward et al., 2012). Eukaryotic algae are restricted to a maximum of ~55 °C. One can observe distinct zones of bright colors and morphologies along a thermal stream that correspond to different species along the thermal gradient. Cyanobacteria dominate geothermal waters, including masses of species of Aphanocapsa, Chroococcus, Cyanothece, and Synechococcus, and filamentous taxa such as Fischerella (formerly Mastigocladus) laminosus, Geitlerinema, and Leptolyngbya (López-Cortéz et al., 2001; Finsinger et al., 2008; Ward et al., 2012). The 16S rRNA sequences and genetic diversity of cyanobacteria (Synechococcus and Oscillatoria-like strains) from Yellowstone hot springs are distinct from those collected on other continents, suggesting that geographical isolation and genetic drift can lead to evolutionary divergence (Papke et al., 2003). Similarly, sequences of Fischerella strains from thermal springs in Costa Rica also differed from isolates collected elsewhere (Finsinger et al., 2008). Cyanobacteria may form mats several cm thick, with extremely high rates of primary production (>10 g C m−2 d−1; Castenholz and Wickstrom, 1975). In one hot spring in Costa Rica (62 °C, pH 7.0), a filamentous cyanobacterium identified as Oscillatoria dominated, while nearby streams with less extreme temperatures (35-36 °C; pH 7.8-8.0) had a greater diversity, with cyanobacteria (Oscillatoria, Phormidium, and Lyngbya) and diatoms (Pinnularia) (Pringle et al., 1993). In other cooler thermal springs (35-50 °C), diatoms (Achnanthidium, Pinnularia) and green algae (Spirogyra, Mougeotia) can proliferate (Stockner, 1967; Castenholz and Wickstrom, 1975). Food web dynamics are not well understood. In very hot systems (>70 °C), bacteriophages may be the only predator of bacteria and cyanobacteria (Breitbart et al., 2004). Other consumers occur in less extreme conditions, but few metazoa tolerate temperatures greater than about 50 °C. Near this limit, invertebrates include ostracods, water mites, and rotifers, but little is known of their dynamics or food webs. Adult beetles and flies are successful in surface water temperatures of 5258 °C (De Jong et al., 2005). Brine flies (Paracoenia, Ephydra) lay eggs in microbial mats of springs in Yellowstone Park within the 30-40 °C range. Both adult and larval stages consume algal and bacterial material, which may enhance primary productivity (Brock, 1967; Brock et al., 1969). Some thermal springs are highly acidic, and their species diversity appears is even more limited. A notable eukaryote, the red alga Cyanidium caldarium, is often the sole photosynthetic organism in acid (pH 2-4) hot springs up to about 55 °C. This enigmatic organism was variously classified as a cyanobacterium, green alga, cryptomonad, or evolutionary link between red and green algae (Seckbach, 1991). Currently it is placed in its own class, Cyanidiophyceae (Rhodophyta), based on pigments, chloroplast structure, and molecular data (Steinmüller et al., 1983; Pueschel, 1990). Other unusual



FIGURE 14 Examples of spring habitats. (A) Thermal (hot) spring-travertine habitat, Mammoth Hot Springs Terraces, Yellowstone National Park (WY); (B) natural acid spring enriched with iron and other heavy metals, Kootenay Paint Pots (BC), with precipitated ochre (ferrous hydroxide) in the littoral zone; (C) karst spring outflow with luxurious growth of filamentous algae, Arch Springs (PA); (D) karst spring rise from an underground cavern, Popo Agie River, Sinks Canyon State Park (WY); and (E) thermally constant, first magnitude spring boil with exceptionally clear water, Volusia Blue Spring (FL). (Photos A–D by J.D. Wehr and photo E by Deborah Donaldson, with permission).

rhodophytes have also been reported, including Cyanidioschyzon merolae and Galdieria sulphuraria (DeLuca and Moretti, 1983). It is unlikely that high temperature or low pH is solely responsible for this peculiar flora, because alkaline hot springs and nonthermal acid springs have very different assemblages.

B Acid Springs There are many nonthermal, highly acid springs and streams that support a characteristic algal flora unlike those in other aquatic environments. While bogs are fairly acidic (4.0-5.0), highly acidic environments are systems with H+ concentrations at least an order of magnitude greater, that is, pH values ≤3.0, that receive acidic inputs from either geological or



anthropogenic sources (Hargreaves et al., 1975). Most have elevated metal concentrations, including Al, Fe, Mn, Pb, Co, Cu, and Zn, which may be near saturation levels, resulting in metal salt precipitates forming along stream margins or on algal colonies. Acid springs with very high Fe concentrations may also have very low dissolved O2, due to Fe(OH)2 and FeO(OH)2 precipitates (Van Everdingen, 1970). The earliest detailed studies of algae in highly acidic systems in North America were conducted in acid mine drainages (AMDs) in Indiana, Kentucky, Ohio, Pennsylvania, and West Virginia (Lackey, 1938; Bennett, 1969; Warner, 1971). These studies, as well as those in the U.K. (Hargreaves et al., 1975) all reveal low species diversity and a remarkable similarity in composition, with Euglena mutabilis the most widespread and often most abundant species, occurring in systems as acidic as pH 1.5. E. mutabilis is also common in naturally acidic streams, such as the Rio Agrio (pH 2.3), Costa Rica (Pringle et al., 1993), and in acidic ponds in the Smoking Hills region of the Northwest Territories (Sheath et al., 1982). The latter site also supports populations of Chlamydomonas acidophila. E. mutabilis is not apparent in natural acid springs in Kootenay Paint Pots (British Columbia), although a few diatoms and green algae occur (Fig. 14B; Wehr and Whitton, 1983). Other common elements of many highly acidic environments include the filamentous green Klebsormidium rivulare, diatoms Eunotia tenella, Pinnularia microstauron, and P. braunii, as well as the gelatinous chrysophyte Gloeochrysis turfosa, although nearly all sites have relatively low biodiversity. Ecological functions, such as biomass accrual rates and enzyme activities of periphyton in AMD streams in Ohio were also significantly reduced, as compared with reference streams in the same region (Smucker and Vis, 2011). Nonetheless, there is evidence that algal species may adapt to these extreme conditions. Isolates of E. mutabilis, Chlamydomonas acidophila, K. rivulare, Gloeochrysis turfosa, and Stichococcus bacillaris from one acid stream (pH 2.6-3.1) tolerated and grew at pH levels less than the lowest measured in their collecting site (Hargreaves and Whitton, 1976a). An acid strain of Klebsormidium rivulare tolerated greater Zn and Cu concentrations in the pH range 3.0-4.0 than at pH ≥ 6.0 (Hargreaves and Whitton, 1976b). It has been suggested that algae may therefore be useful in AMD remediation by immobilizing metals and maintaining sulfate reducing bacteria (Das et al., 2009).

C Karst Springs Karst springs and spring-streams occur within a geologically distinctive topography composed of carbonate rocks, mainly limestone and dolomite and are formed through the dissolution of these rocks by water. Aquatic habitats include a fascinating collection of springs, spring-streams, sinkholes, and subterranean drainages (Fig. 14C–E; Glazier, 2009). Surface habitats possess a remarkable variety of organisms, from bacteria to fish (Minckley, 1992; Ilmonen and Paasivirta, 2005; Dinger et al., 2006). The food webs of these unique ecosystems depend upon algae as their main energy source and as habitat (Mattson et al., 1995; Álvarez and Pardo, 2007). Algal associations in calcareous springs are distinct from those flowing through other geologies, with notably greater diversity of cyanobacteria (Cantonati et al., 2012). Karst springs are especially abundant in north-central Florida (e.g., Fig. 14E) and possess a rich collection of algal assemblages (Whitford, 1956; Mattson et al., 1995; Sheath and Cole, 1990). Many are threatened by nitrogen inputs, causing blooms of invasive algae and reducing biodiversity (Heffernan et al., 2010). Similar threats to karst springs occur in other parts of North America, including the biologically rich Edwards Plateau region in Texas (Bowles and Arsuffi, 1993). Due to high concentrations of carbonates in karst springs and photosynthetic activity of cyanobacteria, eukaryotic algae and bryophytes, CaCO3 precipitation is common, forming tufa around benthic assemblages (Pentecost, 1988; Freytett and Verrecchia, 1998; Golubić et al., 2008). Colonies of Rivularia, Nostoc, and Chaetophora become calcified and can be confused with small pebbles to the untrained eye. Because many karst springs provide highly stable hydrology and chemistry, they often support luxurious growths of filamentous or macroscopic algae (Fig. 14C). In the very clear waters of Cuatro Cienegas in central Mexico, a consortium of cyanobacteria and bacteria form macroscopic freshwater stromatolites with substantial rates of nitrogen fixation (Falcón et al., 2007). DNA sequences and phylogenetic analyses of these stromatolites indicate a high degree of taxa diversity, with many sequences unnamed and not yet present in public databases. Studies conducted thus far in North American karst springs have identified species globally rare algae in fresh waters including the rhodophytes Chroothece mobilis, Sheathia heterocortica, and Polysiphonia subtilissima (Blinn and Prescott, 1976; Sheath and Cole, 1990), chlorophytes Neospongiococcum concentricum and Oocardium stratum (Anderson and Nichols, 1968; Pfiester, 1976), and diatoms Anorthoneis dulcis and Gomphonema montezumense (Czarnecki and Blinn, 1979; Hein, 1991). Fascinating parallels exist in sinkholes deep in Lake Huron (>90 m), karst micro-ecosystems, but with very low light and dissolved O2, and microbial mat assemblages consisting of Eubacteria, Archaea, and purple-colored cyanobacterial mats (Biddanda et al., 2006; Voorhies et al., 2012). There are likely to be more discoveries as algal assemblages in karst systems receive greater study.



VI SUBAERIAL HABITATS Subaerial habitats span a huge range of terrestrial environments, from soils, rock surfaces, plants, and animals to ice and snow. It has been estimated that at least 800 species of green microalgae alone colonize terrestrial habitats (Rindi et al., 2009). These algal assemblages have received less attention that those in lakes or rivers, but their adaptations, taxonomic diversity, and ecological properties offer fascinating challenges. Several reviews and studies summarize this diversity, including Metting (1981), Starks et al. (1981), Cardon et al. (2008), López-Bautista et al. (2007), and Rindi et al. (2009).

A Soils Soils present many conditions that favor algal growth, but also important limitations. Algal cells often colonize soil surfaces, but to minimize desiccation or avoid freezing, they may occur several mm or cm within the soil, where light limitation then becomes an issue. Most soil taxa are microalgae, such as chlorophytes Chlorococcum, Tetracystis, Chlamydomonas, and Chlorella, and cyanobacteria, including Anabaena, Gloeocapsa, Phormidium, Microcoleus, Nostoc, and Scytonema (Hoffmann, 1989; Trainor and Gladych, 1995; Khaybullina et al., 2010; Pentecost and Whitton, 2012). Mats of Microcoleus on damp soils form ropes or braids of filaments that knit together soil particles (Fig. 15A and B). Algal assemblages stabilize desert soil crusts (Johansen, 1993), often with substantial taxonomic and genetic diversity, despite the harsh conditions (Johansen et al., 1981; Lewis and Flechtner, 2004; Cardon et al., 2008). There are also macroscopic soil taxa, including pea-like colonies of the xanthophyte Botrydium (Fig. 15C), gelatinous masses of the cyanobacterium Nostoc (Fig. 15D),

FIGURE 15 Examples of subaerial habitats and algae from soil. (A) Damp soil from a wetland, with a mat of cyanobacteria; (B) bundles of the filamentous cyanobacterium Microcoleus collected from damp wetland soil; (C) spherical colonies of the xanthophyte Botrydium on damp soil; (D) large Nostoc colony collected from a suburban garden; (E) colonies of Protosiphon on soil; (F) green felt of Vaucheria formed on a muddy trail; and (G) Vaucheria prona siphons with reproductive structures, isolated from worm castings (scale = 5 cm for A; B = 100 μm; C and D = 2 cm; E = 1 cm; F = 20 cm; G = 1 mm). (Photos A and B by J.D. Wehr; photos C–E by Chris Carter, with permission; and photos F and G by Craig W. Schneider, with permission).



and expanses of green algae such as Zygogonium, Klebsormidium, and Protosiphon (Fig. 15E). Early studies suggested the soil algal flora was species poor, but recent studies using molecular methods indicate soils are biodiversity “hot-spots” (López-Bautista et al., 2007; Rindi et al., 2009). Algae on soil surfaces are exposed to extremes of light and UV radiation. Cyanobacteria produce mucilage, thick sheaths, or protective pigments, such as scytonemin (Pentecost and Whitton, 2012). The green alga Zygogonium ericetorum produces reduced cytoplasm, thicker walls, and solutes with UV-absorbing capacities (Holzinger et al., 2010). Most species of soil algae are thought to be obligate photoautotrophs (Hoffmann, 1989), but those living below the soil surface may require heterotrophy (Metting, 1981) or exist in a dormant state for extended periods. Viable cells of several green algal taxa were recovered from desiccated soil after 35 years without moisture (Trainor, 1985; Trainor and Gladych, 1995). Subsequent recovery of Protosiphon botryoides after 43 years of desiccation indicates that this alga is especially well adapted to low moisture (Lewis and Trainor, 2012). Zygogonium forms mats up to 6 cm thick on soils (Lynn and Brock, 1969); their high water-holding capacity creates a micro-aquatic environment in which algal flagellates Euglena and Chlamydomonas can persist. Soil algae produce resting cells such as akinetes or zygospores that germinate in response to moisture or temperature cues (Hoffmann, 1989; Agrawal, 2009). The xanthophyte Vaucheria forms a green felt on damp soils, and produces sexual structures in late spring (Gupta and Agrawal, 2007). Propagules (mainly oospores) of Vaucheria from riparian soil are able survive at least 5 months of anoxia (Schneider et al., 2008). Viable propagules can be isolated from earthworm castings (Fig. 15F and G) and hence can survive enteric passage (Schneider and McDevit, 2002). Because earthworms are active in turning over soil, they aid in the dispersal of Vaucheria to other habitats. Soil algae provide important ecological functions, aiding moisture retention, seed germination, and nutrient dynamics (Carson and Brown, 1978; Vazquez et al., 1998). Because nitrogen limits the growth of plants in many terrestrial environments, N2-fixation by cyanobacteria provides ecosystem services for the soils they colonize (Belnap, 2001). Cyanobacterial crusts fix N in short bursts when rain or snowmelt provides sufficient moisture (Belnap, 2002; Housman et al., 2006). They can be the dominant source of N for soils in arid (Belnap, 2002) and arctic ecosystems (Chapin and Bledsoe, 1992).

B Epilithic and Endolithic Habitats The algae colonizing rocks in subaerial habitats may experience even greater stresses with regard to exposure and limited moisture (Gorbushina and Broughton, 2009). On shady or occasionally wet stone walls and walkways one can often observe expanses of green algae (Fig. 16A). Chlorella, Desmococcus, Klebsormidium, and Stichococcus are common in these habitats, but their diversity and abundance are dependent on water availability (Gorbushina, 2007). Diatoms also colonize these habitats; those that are desiccation-resistant are obligate aerial taxa (Johansen, 2010). Prasiola colonizes N-rich (e.g., guano) rocks and walls in aerial and even urban environments (Jackson, 1997; Rindi et al., 1999). The algal flora colonizing cliff faces (Fig. 16B) varies with geology, aspect, exposure, and moisture. While cyanobacteria and green algae are most often observed, xanthophytes (Tribonema sp., Chloridella neglecta, and Ellipsoidion sticchococcoides) and even a dinoflagellate (Glenodinium montanum) have been reported (Camburn, 1983; Gerrath et al., 2000). Epilithic and endolithic taxa are widespread along limestone cliffs of Niagara Escarpment in Ontario, with cyanobacteria and green algal species predominating (Matthes-Sears et al., 1999; Gerrath et al., 2000). Remarkably similar algal assemblages were observed on rock faces in Great Smoky Mountains National Park (TN) and the Lake Superior region (MI, ON); cyanobacteria dominated in drier microhabitats and green algal taxa in wetter sites (Ress and Lowe, 2013). Constructed stone materials equally serve as habitats; algae can be observed on buildings (Fig. 16C), walls, gravestones, and monuments and hasten their deterioration (Ortega-Morales et al., 2000; Pentecost and Whitton, 2012). A number of algal taxa colonize the interstitial spaces within crystalline rocks in arid or semiarid regions (Fig. 16D and E; Bell, 1993; Johansen, 1993). Endolithic algal assemblages colonizing sandstone in more mesic regions of the Colorado Plateau include a number of coccoid and sarcinoid genera such as Chlorococcum, Borodinella, Chlorosarcinopsis, and Tetracystis. Cyanobacteria (e.g., Chroococcidiopsis, Gloeocapsa) dominate in more xeric locations (Bell et al., 1988). Biomass and primary production within these rocks were low, but contributed 5-10% of total autotrophic biomass and 2080% of primary production in this system (Bell and Sommerfeld, 1987). Algae colonize caves in limestone regions where fractures or underground streams create small pockets or large caverns (Albertano, 2012). With limited light, cyanobacteria tend to dominate and rely on phycobilin pigments and thylakoid rearrangement (Albertano, 2012; Mulec et al., 2008; Baulina, 2012). While many caves can be relatively dim, unproductive environments, algal floras dominated by cyanobacteria develop where surfaces are open to the sunlight (fissures, mouths) or in show caverns where artificial lighting has been added (Claus, 1962; Round, 1981; Gurnee, 1994). Species diversity is typically low. Timpanogos Cave (UT) and Seneca Cavern (OH) each support an algal flora of fewer than 30 species (St. Clair and Rushforth, 1976; Dayner and Johansen, 1991).



FIGURE 16 Examples of subaerial habitats and algae on and in rock. (A) Green algal assemblage on a limestone walkway; (B) mixture of cyanobacteria and green algae on a cliff wall seepage in Glacier National Park (MT); (C) green algal growth on a sandstone wall and staircase in NYC; (D) macroscopic view of Coconino sandstone sample collected from Arizona semidesert habitat, broken open to expose of cryptoendolithic algae layer, scale bar = 1 cm; and (E) close-up view of cryptoendolithic algae from same habitat and sandstone type as (D), most colonies consist of the cyanobacterial genus Chroococcidiopsis, scale bar = 2 mm. (Photos A and C by J.D. Wehr; photo B by Janet R. Stein with permission; and photos D and E by Todd Huspeni and Robert Bell, with permission).

C Plants and Animals Terrestrial plants also serve as habitats for algae, including their surfaces, depressions, cavities, and internal tissues. Algae colonize the cup or phytotelmata of the pitcher plant Sarracenia purpurea at densities up to 105 cells mL−1, where there is a succession of taxa. Coccoid and flagellated green algae, chrysophytes, and the xanthophyte Bumilleriopsis sp. can predominate in younger pitchers, with filamentous green algae in older cups (Gebühr et al., 2006). Algae also occupy bromeliad cups, habitats that differ in pH and dissolved O2 levels depending on the plant species (Laessle, 1961). Algal densities range up 104 cells mL−1, accounting for up to 30% of carbon in these microbial food webs (Brouard et al., 2011). In temperate



climates, coccoid green algae such as Desmococcus olivaceus (syn. Pleurococcus vulgaris) and Apataococus spp. colonize tree bark (e.g., Fig. 17A, Hoffmann, 1989), appearing as a green tint or powdery green patches. In humid climates, branched filamentous green algae (members of Trentepohliales) colonize the bark, leaves, and even fruits of plants (Chapman, 1984; Thompson and Wujek, 1997). Assemblages appear as rusty-colored (Trentepohlia) or green (Rosenvingiella) fuzz, which can completely cover plant surfaces or wooden fences (Fig. 17B and C). Nostoc colonies live among the leaves of terrestrial bryophytes, including the liverwort Porella avicularisa in coastal California to British Columbia (Dalton and Chatfield, 1985) and feather mosses Pleurozium schreberi and Hylocomium splendens in boreal forests (Zackrisson et al., 2004; Ininbergs et al., 2011). P. schreberi secretes species-specific chemoattractants when N-limited, which guide Nostoc motile hormogonia toward them (Bay et al., 2013). Root tissues of some

FIGURE 17 Examples of subaerial habitats and algae on and in plants. (A) Thin film of green algae (Apatococcus and/or Desmococcus sp.) on tree bark (NY); (B) rusty covering of Trentepohlia on a tree trunk in tropics (Costa Rica) and (C) on a wooden fence (CA); (D) cross section of the root of Gunnera manicata showing patches of cyanobacteria (arrows); and (E) rusty-brown patches of Cephaleuros, parasitizing leaves in the tropics (CR). (Photo A by J.D. Wehr; photos B and E by Juan M. Lopez-Bautista, with permission; and photos C and D by Janet R. Stein, with permission).



cycads (Cycas, Encephalartos, and Macrozamia) and the wetland plant Gunnera manicata are colonized by several cyanobacteria, including Anabaena, Calothrix, and Nostoc (Huang and Grobbelaar, 1989). When sectioned, cyanobacteria appear a healthy blue-green color despite being embedded within the plant tissue (Fig. 17D). Cyanobacterial associations in hornworts are apparently ubiquitous (Adams et al., 2012). For example, Anthoceros is a habitat for endosymbiotic Nostoc spp., which fix N inside the host (Rodgers and Stuart, 1977). In Costa Rican rainforests, Nothoceros superbus also harbors abundant colonies of Nostoc (Villareal et al., 2007). Intimate algal or cyanobacterial symbioses form with fungi to form lichens. Most lichens are terrestrial and well adapted to xeric conditions. But several colonize aquatic habitats, including mountain streams (Glavich, 2009) and rocky lake margins (Wurzbacher et al., 2010). While the most common alga in terrestrial lichens is the genus Trebouxia, this alga is rare in aquatic systems. The cyanobacteria Calothrix (in Lichina), Nostoc (in Pyrenocollema), Stigonema (in Ephebe), and the green algae Stichococcus (in Staurothele), and Dilabifilum (in Verrucaria) are common in aquatic lichens (Hawksworth, 2000). Cephaleuros is a parasite (Fig. 17E), the cause of red rust disease in more than 40 families of flowering plants (Thompson and Wujek, 1997; López-Bautista et al., 2006). There are examples of terrestrial animals as habitats for algae. An unusual example is the hair of two- and three-toed sloths, which in which the red alga Rufusia (Wujek and Timpano, 1986) and cyanobacterium Oscillatoria pilicola (Wujek and Lincoln, 1988) have been recorded. The hairs have grooves that absorb water, with algal cells rendering the hair green. A study of sloth samples using partial 18S rRNA gene sequences has since uncovered a fairly diverse eukaryotic algal assemblage, with green algal taxa predominating (Suutari et al., 2010). Between 6 and 22 distinct taxa from the green algal class Ulvophyceae were identified from sloth hair samples collected in Panama and Costa Rica, often with Trichophilus welckeri, a species first described in 1887 (Hoffmann, 1989).

D Snow and Ice Aristotle apparently observed red snow more than two millennia ago (Kol, 1968), and people who have hiked in alpine regions are familiar with this phenomenon (Fig. 18A). It is most often caused by the green alga Chloromonas (syn. Chlamydomonas) nivalis. The red color is the result of an accumulation of carotenoids, mainly astaxanthin, in resting cells (Bidigare et al., 1993). But not all assemblages are red; diatoms and cyanobacteria are also observed where orange, brown, and green patches are also seen, depending on the species, pigments, sunlight, and pH (Stein and Amundsen, 1967; Hoham and Blinn, 1979). Cells aggregate near the surface, and as successive snowfalls accumulate, layers or bands of pigmented algal communities can be seen in vertical cuts through a snow bank (Fig. 18B; Hoham and Mullet, 1977). The composition of snow assemblages on the Alaska Range vary along an altitudinal gradient, with greater densities below 1600 m (Takeuchi, 2001). Ancylonema nordenskioeldii was the most abundant alga on lower portions of the glacier, while Chloromonas nivalis was the predominant alga at upper elevations. While motile stages do occur in snow, most cells occur either as thick-walled resting zygotes or asexual hypnospores (Stein and Amundsen, 1967). This cryophilic flora includes

FIGURE 18 Examples of snow and ice algae. (A) Green snow in several layers below the surface, from 1400 m elevation at Cayuse Pass, Mt. Rainier National Park (WA), caused mainly by Chloromonas brevispina with some Chloromonas nivalis and (B) red snow at 2040 m elevation near Frozen Lake, Mt. Rainier National Park (WA), caused by Chlamydomonas nivalis. (Photos by Ronald Hoham, with permission).



other algal flagellates, such as Chloromonas brevispina, Carteria nivale, Scotiella cryophila, and Chromulina chionophila, and nonflagellated green algae such as Raphidonema nivale and Stichococcus spp. (Kol, 1968; Hoham, 1975; Hoham and Blinn, 1979). Snow algae exhibit measurable but low photosynthesis rates, many of which reach their maximum at 10 °C, although some peak near freezing and decline at higher temperatures (Mosser et al., 1977). Algal cells on snow surfaces at high altitude experience extreme levels of solar radiation, and many employ carotenoid pigments as a photoprotective function (Bidigare et al., 1993). Experiments with algae from Tioga Pass in the Sierra Nevada demonstrated that UV radiation inhibited photosynthesis in green snow by 85%, but only 25% in red snow (Thomas and Duval, 1995).

ACKNOWLEDGMENTS Thanks are due to Dr. Alan Steinman (Grand Valley State University) and Dr. Walter Dodds (Kansas State University) for very helpful reviews and advice on relevant literature. We also thank Robert Bell, Dean W. Blinn, Chris Carter, Andrew F. Casper, Don Chamberlain, Todd A. Crowl, Deborah A. Donaldson, Ronald W. Hoham, Todd Huspeni, Louise A. Lewis, Juan M. Lopez-Bautista, Louis J. Maher Jr, Craig W. Schneider, Janet R. Stein, James H. Thorp, Kam Truhn, Yuuji Tsukii, and Warwick F. Vincent for the use of photos used in this chapter.

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