Managed Flooding for Riparian Ecosystem Restoration

Managed flooding reorganizes riparian forest ecosystems along the middle Rio Grande in New Mexico

Manuel C. Molles Jr., Clifford S. Crawford, Lisa M. Ellis, H. Maurice Valet!, and Clifford N. Dahm

AcenturieS-Old Indian pueblo named Isleta stands on a small rise of land west of the Rio

Grande in central New Mexico. Its name, which means "island" in Span­ish, reflects the fact that the pueblo was on an island in the middle of the Rio Grande when it was first seen by Spanish explorers (Pearce 1965), The name Isleta comes from the time be­fore flood control on the Rio Grande, when spring flood waters roared down from the San Juan Mountains of Southern Colorado, flooding the middle Rio Grande Valley of central New Mexico (Figure 1). During some years, high spring runoff would ex­tensively flood the middle Rio Grande Valley, sending most valley residents fleeing to surrounding high ground until the flood waters receded. How­ever, these floods rarely touched Isleta Pueblo, only transforming it into the island its name suggests. While the residents of other settle­ments huddled in tent villages on the surrounding mesas, the life of Isleta's residents went on as usual. Today, dams and levees built on the Rio ------~----

Manuel C. MoUes Jr. (molles@sevilleta. unm.edu) is a professor, Clifford S. Crawford (ccbosque@unm.edu) is a pro­fessor emeritus, Lisa M. Ellis (lellis@ sevilleta.unm.edu) is a graduate student, and Clifford N. Dahm (cdahm@sevilleta. unm.edu) is a professor in the Department of Biology, University of New Mexico, Albuquerque, NM 87131. H. Maurice Valett (rnvalett@vt.edu}isanassistantpro­fessor in the Department of Biology, Vir­ginia Polytechnic Institute and State Uni­versity, Blacksburg, VA 24061. © 1998 American Institute of Biological Sciences.

September 1998

Out of this search

for general ecological

principles will come

a better understanding

of the structure

and function of

riparian ecosystems

Grande have reduced the frequency and intensity of flooding; as a result, Isleta Pueblo is surrounded not by flood waters each spring but by alfalfa fields and suburban development.

The flood-pulse concept IJ unk et al. 1989, Bayley 1995) emphasizes that water-land interactions create and maintain river-floodplain eco­systems as some of the most produc­tive and diverse ecosystems in the world. However, river regulation has largely eliminated the flood pulse from most large rivers. Postel et al. (1996) estimated that humans now use a little over half of accessible river runoff worldwide. Fragmenta­tion of river channels by dams, inter basin diversion, and irrigation strongly or moderately affect 77% of the total discharge of the 139 largest rivers in the northern third of the earth (Dynesius and Nilsson 1994). Benke (1990) noted that, ex­cept for the Yellowstone River in Montana, all large rivers in the 48 contiguous United States have been severely altered for flood control,

hydropower, or navigation. He also estimated that of the 5,200,000 km of streams and rivers in the contigu­ous 48 states, only approximately 2% are classified as high quality. Moreover, only 42 free-flowing river segments longer than 200 km in length remain. There are currently 75,000 dams on the streams and rivers of the United States (Meyer 1996), and large dams are being com­pleted at an estimated rate of 500 per year worldwide (Covich 1993). Thus, the extensive disconnection of rivers from their floodplains repre­sents global-scale ecological change, the consequences of which we do not understand.

Studies of the effects of modified river flow on riparian ecosystem structure and function require long­term flow records. The flow record for the middle Rio Grande is excep­tionally long. The first stream gauge established by the US Geological Survey was constructed on the Rio Grande at Embudo, New Mexico, in 1889. Since its construction, the Embudo gauge has been measuring the hydrologic regime of the Rio Grande. The most salient feature of that regime is a consistent annual peak in flow at the end of Mayor the beginning of June (Slack et al. 1993). Although flows still peak at the end of May, dams on the Rio Grande have eliminated most flooding. His­torically, floods beginning in May would sometimes continue until late July. This spring flood pulse made the middle Rio Grande Valley a mo­saic of river channels, marshes, wet meadows, and riparian forests of

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Figure 1. A map of the middle Rio Grande Val­ley in New Mexico showing the location of the study area at Bosque Del Apache National Wildlife Refuge.

various ages. T aday, however, the middle Rio Grande no long-er meanders freely across its valley but is constrained to flow within its levees. Most riparian for-ests are now con-fined to the area be-tween the levees, and most of the wetlands in the middle Rio Grande have disap-peared (Crawford et al. 1993, 1996a).

What are the eco­logical consequences of eliminating the

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annual flood pulse in the middle Rio Grande Valley? One of the most ob­vious consequences has been a dra­matic reduction in the germination and establishment of native cotton­woods and willows (Howe and Knopf 1991). The riparian forest along the middle Rio Grande in central New Mexico is the most extensive cotton­wood-willow forest left in the south-

Rio Puerco

Rio Salado

Cochiti Reservoir

Jemez River

Cochiti Reservoi

Albuquerque

Bosque del ApacheNWR

western United States (Crawford et a1. 1993). However, this extensive forest is largely an ecological legacy of past flooding. These stands may be rapidly senescing, and because of flood control, new stands of cotton­wood are not being established. In addition, a stabilized river flow ap­pears to favor invasion by non-na­tive tree species, such as saltcedar,

Figure 2. A patch of riparian forest at Bosque Del Apache National Wildlife Refuge in ] 991. This patch, which had not flooded for over 50 years, accumulated large amounts of woody debris. The absence of a flood pulse may contribute to such accumulations.

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T amarix ramosissima, and Russian olive, Elaeagnusangustifolia (Camp­bell and Dick-Peddie 1964, Crawford et al. 1996a).

Students from the Department of Biology at the University of New Mexico have been monitoring sev­eral stands of riparian forest, or "bosque," as riparian forests along the Rio Grande are known, for ap­proximately ten years. Monitoring of these riparian forests was prompt­ed by the concern that without long­term data sets, these natural areas could not be managed properly. The riparian forest monitored by these students was established mainly dur­ing the last large flood on the middle Rio Grande, which took place in 1941-1942. Now, however, most of the forest along this reach no longer experiences overbank flooding. In the absence of flooding and with the invasion of non-native trees, these tracts of riparian forest have become tangled with woody debris (Figure 2). The organic matter that has accu­mulated on the floor of riparian for­ests along the middle Rio Grande now averages over 50 x 106 g/ha in some areas (MoUes et al. 1996), which increases fire danger (Stuever 1997) and immobilizes substantial quantities of plant nutrients.

Monitoring has produced several data sets , including leaf fall, which has been used as an index of forest production. Autumn leaf fall has been measured in three sections of a 760 m long segment of riparian forest within the Rio Grande Nature Cen­ter in Albuquerque, New Mexico. From 1989 to 1996, leaHall dropped steadily, declining by over 40% in just eight years. This record of leaf fall quantifies what had previously only been suspected: The bosque within Albuquerque is senescing rap­idly (Figure 3). This finding gives a sense of urgency to research on ri­parian restoration along the middle Rio Grande.

Little is known about the poten­tial importance of flooding for main­taining established riparian forest. To investigate how the absence of a flood pulse on the middle Rio Grande has affected the structure of micro­bial and animal populations on the forest floor and how the flood pulse affects the rates of ecosystem pro­cesses such as forest floor respira-

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tion and decomposition of plant lit­ter, we created a series of managed floods in a riparian forest at Bosque del Apache National Wildlife Ref­uge near San Antonio, New Mexico. A major objective of our research is to evaluate the potential of managed floods as a tool for riparian restora­tion by measuring key population, community, and ecosystem responses to the reestablishment of flooding. In this article, we present some of our findings that bear on this ap­proach to riparian restoration.

Restoration and managed flooding

Figure 4 presents a conceptual model of the present status of and restora­tion efforts in the Rio Grande bosque. The present riparian ecosystem is a complex mosaic of forest patches dominated by either non-native veg­etation, represented in Figure 4 by saltcedar, or native cottonwoods. These patches can be further subdi­vided into those that are disconnected from the river and are subject to long interflood intervals, and those that remain connected to the river and are subject to short interflood inter­vals. Current restoration efforts in­valve two main approaches. The first is aimed at converting non-native forests to native forest by uprooting saltcedar and establishing cotton­wood forests. This approach is indi­cated by the arrows labeled "1" in Figure 4. The second approach is to use managed flooding to reconnect disconnected riparian forests to the Rio Grande-that is, to reduce inter­flood intervals. It is this restoration approach (indicated by the arrow labeled "2" in Figure 4) that we explore in our studies.

Two study sites were established in 1991 in mixed cottonwood forest at Bosque del Apache National Wild­life Refuge, elevation 1400 m, ap­proximately 5 km south of San An­tonio, New Mexico. The two sites, which lie outside the river levee, have been isolated from flooding by the Rio Grande for more than 50 years. One site was designated as a control site and the other as the experimen­tal flood site. Although we flooded approximately 10 ha of forest, eco­logical monitoring was organized around a 3.1 ha, 200 m diameter

September 1998

Figure 3. Annual leaf fall at the Rio Grande Na-ture Center in Albuquer-que, New Mexico, from 1989 to 1996. Leaf fall declined by approxi-mately 40% over that period. The absence of a flood pulse in the mid-die Rio Grande may reduce riparian forest leaf production.

mammal-trapping web. In 1994, we es­tablished two addi­tionalstudy sites with-

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in the levees at Bosque del Apache. One of these additional sites seldom floods and so served as a second dry control site. The other site has con­tinued to flood frequently because of its topography and position within the levees and was therefore desig­nated as the natural flood site.

The canopy at all four study sites was dominated by Rio Grande cot­tonwood, Populus deltoides ssp. wislizenii. ranging from 8 to 15 min height. The forest subcanopy con­sisted of Goodding willow, Salix gooddingii, and saltcedar. The domi­nant understory shrubs were seep­willow, Baccharis glutinosa. New Mexico olive, Forestiera neomexi­cana, and Russian olive.

We collected baseline data for two years and then flooded the experi­mental flood site for approximately one month during each of the fol­lowing three years (Figure 5). These managed floods were designed to simulate the conditions of a low­energy flood near the edge or in the backwater of a major flood and to saturate riparian soils and soak or­ganic matter on the forest floor. Floods, which took place from mid­May to mid-June of 1993, 1994, and

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1995, were timed to coincide with the historic flood peak of May 31 at the US Geological Survey's Embudo gauging station on the upper Rio Grande (Slack et al. 1993). Flood­water was diverted from a nearby irrigation canal that carried a mix­ture of water diverted directly from the Rio Grande, irrigation return flows, and groundwater drainage ac­cumulated in the nearby Rio Grande Low Flow Channel. The level of water diverted onto the experimental flood site was controlled by preexisting berms and by water-control gates constructed by personnel of the Bosque del Apache National Wild­life Refuge at the inflow and outflow of the experimental flood site. Be­cause the topography is uneven, water depth during managed floods ranged from approximately 20 cm to over 200 cm, with an average flood­water depth of approximately 50 cm.

An ecosystem reorganization model for riparian restoration

Although Figure 4 depicts the rees­tablishment of the flood pulse to restore native forest as a direct pro­cess, reestablishing the flood pulse

Dominant Riparian Vegetation

Saltcedar Cottonwood

connected connected

Short non-native native I

Figure 4. Riparian forest resto­ration in the middle Rio Gr'ande Valley. Approach 1 converts non-native forest into native cottonwood forest by uproot­ing saltcedar and establishing cottonwood through pole planting or wet-soils manage­ment to favor cottonwood seed germination. Approach 2 rees­tablishes the flood pulse, re­connecting riparian forest to

the river and consequently re­ducing the interflood interval.

Interflood Interval disconnected disconnected

non-native native Dz I Long

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to a riparian forest long disconnected from its river will initiate a distinc­tive ecological phase. Based on ob­servations made during our research, we propose that flooding previously isolated riparian forests will initiate significant reorganization of the ri­parian ecosystem, a process much like that proposed by Bormann and Likens (1979) for hardwood forest ecosystems undergoing succession following clearcutting. We suggest that following reestablishment of flooding along the Rio Grande, this process of reorganization will even­tually return the riparian forest to a position similar to its historic state (Crawford et al. 1996b). We pro­pose that the ecological response to the restoration of flooding will in­volve three phases: the initial discon­nected phase, represented in this study by the two dry control sites; a reorganization phase initiated by managed floods and represented in this study by the experimental flood site; and a steady-state phase that will approximate conditions prior to flood control, represented in this study by the natural flood site.

The structure and function of the riparian forest should differ substan­tially during the disconnected, reor­ganization, and steady-state phases. For instance, Figure 6 shows our model for changes in rates of forest­floor respiration during these phases. During the disconnected phase, for­est-floor respiration will be limited by moisture availability. Indeed, as a consequence of moisture limitation in the arid climate of central New Mexico, forest-floor respiration should be orders of magnitude lower during the disconnected phase than during the other two phases. We predict that in mesic regions, forest­floor respiration should be more simi­lar during the three phases.

The model predicts that forest­floor respiration will be highest dur­ing the initial stages of the reorgani­zation phase, when organic matter pools are at their maximum and the pool of relatively labile organic mat~ ter is still large. As reorganization proceeds, annual flooding will con~ tinue to produce pulses of intense forest-floor respiration, but the height of the respiratory peak should be progressively damped as decom­position reduces the quantity and

752

quality of organic matter on the for­est floor. Eventually, when the pool of organic matter consists princi­pally of the previous season's litterfall, the respiratory peaks will level off and the system will enter the steady-state phase.

Ecological responses to managed flooding

The first floodwaters diverted onto the experimental flood site crossed land that had not been subject to the flood pulse of the Rio Grande for over half a century. The rising wa­ters of this managed flood induced immediate responses by forest-floor animals. The forest floor at the lead­ing edge of the floodwaters came alive with hopping crickets and run­ning spiders. However, our measure­ments revealed that some of the most dramatic responses to managed flooding involved unseen microbial populations and chemical and physi­cal processes (Ellis et a1. 1996). A broad range of ecological responses to the managed flooding conducted in this study is summarized in Table 1.

Population and community responses to flooding. Microbial populations and activity were generally increased by flooding (Table I). Indications of increased microbial activity at the experimental flood site compared to the control site included increased abundance of soil bacteria, fungi, and cellulose decomposers. Both the root lengths colonized by mycor­rhizal fungi and the mycorrhizal in­oculum potential were also generally higher at the experimental flood site than at the control site. Dehydroge­nase activity, which indicates soil biological activity, was more than twice as high at the experimental flood site.

Flooding also restructured the for­est-floor arthropod community. Dur­ing the course of the study, 120 spe­cies of forest-floor arthropods were collected at the control site and 116 species were collected at the experi­mental flood site. Managed flooding did not appear to affect the number of forest-floor arthropod species, but it did alter the relative abundance of species. Although flooding reduced both the species richness and the abundance of ants at the experimen-

tal flood site, at least one species of arboreal ant, Crematogaster cerasi, thrived in the presence of managed flooding. By contrast, floods reduced the abundance of ground-nesting ant species, such as Monomorium mini~ mum. Flooding also reduced the abundance of two non-native terres­trial isopods, Armadillidium vulgare and Porcellio laevis. Meanwhile, populations of the native floodplain cricket, Gryllus alogus, increased with managed flooding.

In contrast to the responses by microbial and arthropod popula­tions, three years of flooding had no measurable effect on small-mammal populations (Ellis et a!. 1997). The dominant small mammal, the white­footed mouse, Peromyscus leucopus, was abundant at both the dry con­trol sites and at the experimental and natural flood sites. Flooding did not affect the abundance of P. leucopus or its reproductive condition. Traps set in cottonwood trees during flood­ing showed that P.leucopus remained in the forest throughout the man­aged floods. This adaptability to managed flooding is not surprising because P. leucopus is common in floodplain forests throughout North America (Ruffer 1961, Blem and Blem 1975). This highly arboreal mouse escapes floodwaters simply by climbing trees.

Ecosystem responses to managed flooding. Large quantities of dis­solved and particulate organic mat­ter in the flooded forest created a very high biological oxygen demand, approaching that of raw sewage (Lieurance et a!. 1994). The flood­waters entering the forest held ap­proximately 8 mg/L of dissolved oxygen (DO) but were rapidly de­pleted of DO by high oxygen de­mand. DO concentrations at 15 em above the submerged litter layer fell to undetectable levels throughout the managed floods. By contrast, at the natural flood site, DO concentra­tions never fell below 4 mg/L within a few centimeters of the forest floor.

We used respiration chambers placed on the forest floor to measure forest-floor respiration during two managed floods and one natural flood. During the second managed flood, respiratory rates at the experi­mental flood site were 300 times

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Table 1. Several population, community, and ecosystem variables at control (C) and experimental flood (EF) sites.

1993 1994 1995 Response' C EF C EF C EF

Total soil bacteria (x 107 colonieslg soil) 431~ (±6)C 401 (±4) 371 (f6) 591 (±4) 391 (±4) 662 (±6) Total soil fungi (x 103/g soil) 673 (±8) 341 (fs) 45 2 (f3) 693 (±S) 36 1,l (±4) 823 (±9) Cellulose decomposers (x 105 colonies/g 20 1 (±3) 52 2 (fs) 482 (±4) 93] (±6) 46 2 (±S) 943 (±7)

soil) Dehydrogenase ().lg • g-1 , h-1) 4.19 1 (fO.45) 9.262 (±0.62) 3.751 (±OAO) 9.42 (fO.66) Mycorrhizal colonization (% root length) lll(±4) 26l (±2) 48 3,4 (±4) 494,5 (f7) 403.4 (f6) 56 5 (is) Mycorrhizal inoculum potential 321 (±4) 50' (±4) 35 1 (±4) 562 (±9) 33 1 (±9) 56 2 (is)

(% root infection) Ants (total in 30 traps in June and August) 348 18 192 9 95 10 Isopods (total in 30 traps in June 970 320 407 49 2,335 405

and August) Crickets (total in 30 traps in June 2 5 2 84 4 9

and August) Mice (individualslha in August) 22.021 (f3.09) 25.98 1 (±3.3s) 16.741 (±l.72) 13.661 (±lAS) Log decomposition (% mass remaining) 96.06 1 (±lA8) 91.76 1 (±l.ll) 95.781 (f1.29) 84.792d (±S.07) leaf decomposition (g dry mass remaining) 2.781 (fO.07) 2.182 (±0.03) 3.321 (±0.10) 1.782 (±0.06) 3.271 (±0.06) 1.982 (±0.02)

'Statistical comparisons for microbial and fungal data are across all sites and years within each row; comparisons for mice and log and leaf decomposition are control versus experimental flood for each year only. bDifferenrly numbered superscripts indicate significant differences (F < 0.01). <Values in parentheses are standard errors. dp = 0.061

higher than at the unflooded control site. During the third managed flood, forest-floor respiration was nearly 600 times higher at the experimental flood site than at the control site. The extremely high rates of respira­tion in the flooded forest appear to have involved many metabolic path­ways, including sulfate reduction. During all three experimental floods, the flooded forest was filled with the smell of many sulfurous compounds. However, sulfurous smells were pro­gressively reduced with each man­aged flood, which suggests ecosys­tem reorganization. Respiration rates measured at the natural flood site were approximately one-tenth the rates measured at the experimental flood site.

Flooding increased the rate of mass loss by leaves and logs (Molles et aJ. 1995, Ellis et a!. 1996). By the sec­ond and third managed floods, mass loss by leaves at the experimental flood site was 125% higher than at the control site. However, the rates of leaf decomposition were almost identical at the experimental flood and natural flood sites. Flooding probably increases the rate of leaf mass loss through leaching of soluble substances and by increasing bacte­rial and fungal activity (Webster and Benfield 1986, Biirlocher 1992).

Log decay rates were calculated using the exponential model Yt = yoe-kt, where Yo is the initial log mass,

September 1998

Yt is the log mass at time t, and k is the decay rate constant. At the ex­perimental flood site, the average rate of log decay, k, was 0.065 yo', whereas at the control site it was 0.010)'1. These decay constants yield log half-lives of 10.6 years at the experimental flood site and 69.3 years at the control site. Although the three experimental floods did not significantly reduce the quantity of organic matter at the experimen­tal flood site, the state of decay of woody debris was more advanced at this site than at the control site (Ellis et al. 1996, Molles et al. 1996).

Patterns of ecosystem reorganization

The population, community, and ecosystem responses we identified indicate that managed flooding initi­ated a process of ecosystem reorga­nization at the experimental flood site. Figure 7 pairs the conceptual model from Figure 6 with measure­ments of forest-floor respiration at the control, experimental, and natu­ral flood sites. As predicted by the model, rates of respiration were two orders of magnitude higher during the experimental flood than on the dry forest floor. Contrary to the model, however, forest-floor respi­ration was 50% higher during the second experimental flood than it was during the first experimental

flood. This increase may be the re­sult of less anoxia during the second flood, which produced higher rates of organic matter processing. Also as predicted by the model, forest-floor respiration at the natural flood site was intermediate between the rates observed at the control and experi­mental flood sites.

The qualitative agreement be­tween the ecosystem reorganization model and measured rates of forest­floor respiration indicates that res­toration of flooding to disconnected riparian forests initiates a distinctive reorganization phase during the tran­sition to full restoration. However, we predict that the amount of time required for reorganization will vary a great deal from one ecosystem com­ponent to another. As a consequence, some aspects of restoration will be delayed longer than others. Although long-term study and continued flood­ing of study sites will be necessary to document variation in the tempo of reorganization, it is possible to make some preliminary predictions.

The results of the three managed floods indicate that the time required for restoration of ecosystem compo­nents will range from less than one year to decades. For instance, the minimal response of small-mammal populations to the three managed floods suggests that restoration of flooding will not lead to a reorgani­zation of these populations. By con-

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trast, populations of decomposer fungi, mycorrhizal fungi, and sur­face-active arthropods, such as ants and crickets, all showed substantial responses to the experimental floods and are therefore subject to reorga­nization. Both mycorrhizal and de­composer fungi doubled their activ­ity during the course of each flood at the experimental flood site but re­turned to levels comparable to those at the control site within months. Thus, reorganization of fungal popu­lations appears to take less than a year. Surface-active arthropods showed a greater response to flood­ing than fungi, but the response ap­pears to have built up over the course of the three floods and not to have reached steady state. Consequently, reorganization of surface-active ar­thropod populations will likely take several additional years. By contrast, leaf decomposition was virtually identical at the experimental flood and natural flood sites by the third managed flood, indicating that reor­ganization of this process was com­plete.

Reorganization of forest- floor or­ganic matter should take much longer. The orders-of-magnitude in­crease in forest-floor respiration in response to experimental flooding indicates a dramatic response by the forest-floor community. However, even -with this high rate of respira­tion during flooding, three managed floods did not decrease the amount of forest-floor organic matter at the experimental flood site (MoUes et al. 1996). The predicted half-life of 10.6 years for logs at the experimental flood site suggests that reorganiza­tion of forest-floor litter may take a decade or more.

Implications for management, policy, and research

The widespread modification of river and riparian ecosystems creates an urgent need to better understand the ecological effects of isolating ripar­ian ecosystems from rivers and to develop methods to restore or better manage these threatened ecosystems. A growing scientific consensus iden­tifies river and riparian restoration as priority areas for future research. Naiman et al. (1995) prioritized six research areas in freshwater ecol-

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ogy. The first priority was to develop principles to guide restoration of aquatic ecosystems and to develop methods to assess the success of res­toration, the second was to deter­mine principles for maintaining biodiversity, and the third was to explore the effects of modified hy­drologic flow patterns on ecological processes. These three priorities are addressed by the effects of our man­aged floods on the composition of forest-floor communities and on rates of organic matter processing and by the ecosystem reorganization model for riparian restoration that we have proposed to explain these effects.

Restoration of river and riparian ecosystems has become a topic of intense interest worldwide. The man­aged flood on the Colorado River in the southwestern US in the spring of 1996 focused scientific and public attention on methods for managing and restoring river and riparian eco­systems through managed flooding (Collier et al. 1997; Schmidt et al. 1998). Reestablishing flooding on a river whose flow has been regulated for decades is now seen as a potential means to restore riverine fluvial geo­morphology and to improve the chances for survival of threatened and endangered species in a riverine landscape. Another major restora­tion project is directed at the channelized and regulated Kissimmee River in South Florida. This project, which is scheduled to begin in 1998, is designed to restore 70 km of river channel and 11,000 ha of wetlands over the next 15 years (Dahm et al. 1995, Cummins and Dahm 1995, Toth et al. 1998).

Ambitious projects such as these represent historic initiatives in eco­system restoration; however, they are a small part of the challenges that remain to restore degraded rivers and riparian zones throughout the world. Our research has yielded some of the first quantification of riparian ecosystem response to managed flooding. The observed rates of re­sponse and the ecosystem reorgani­zation model for riparian restora­tion provide estimates of the potential trajectory and time required for ri­parian restoration with managed flooding.

Nowhere are anthropogenic im­pacts on river ecosystems greater than

in the arid and semi-arid regions of the world. Water is the lifeblood of these regions, which constitute ap­proximately 33% of the land surface worldwide. Increasing human de­mand for fresh water in arid regions leads to inevitable conflicts between the water needs of human popula­tions and those of native ecosystems. The mismatch between water avail­ability and water demand has placed most rivers and riparian zones of arid lands in peril. Based on historic habitat losses and increasing de­mands for water in the region, Christensen et al. (1996) describe all riparian forests in the United States as threatened ecosystems and the ri­parian forests of New Mexico, Ari­zona, and California as endangered.

Research to determine the basic principles for protecting and restor­ing these endangered ecosystems is, therefore, urgently needed. Our own studies have examined a broad spec­trum of responses to managed flood­ing by a riparian ecosystem ina semi­arid region. Some of the results of this research have already been in­corporated into policy and manage­ment recommendations for the Rio Grande (Crawford et a1. 1993, 1996b). For instance, water manag­ers along the Rio Grande are now considering a program of managed flooding, especially during years of higher water availability, to safe­guard the health of the riparian for­est. With this same goal in mind, academic scientists and water man­agers are conducting cooperative studies to quantify the potential water costs of managed flooding of riparian forests along the middle Rio Grande in New Mexico. The results of this work along the Rio Grande may have management implications for large, regulated rivers in other semi-arid regions.

There has been a growing under­standing of the functional role of riparian zones (e.g. Gregory et al. 1991, Stanford and Ward 1993, Bayley 1995), which are now recog­nized as areas of high primary pro­duction, exceptional biodiversity, and rapid biogeochemical cycling. Our research applies an ecosystem perspective to riparian zones, simi­lar to that emphasized by Gregory et al. (1991) and Beneala (1993), in whieh hydrologic exchange integrates river-

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Figure 5. The experimental flood site at Bosque Del Apache National Wildlife Rcf~ uge in flood in 1994. Managed floods on this site flooded approximately 10 ha of riparian forest that had been disconnected from the Rio Grande for over 50 years.

ine and catchment systems and strongly influences biogeochemical processes and populations in ripar­ian zones. Because flood control has limited the connections between riv­ers and riparian ecosystems, we are particularly interested in determin­ing how variation in the interflood interval affects ecosystem structure and function. In most cases, reestab­lishing the full spectrum of historical river-riparian interactions is not a feasible management opt ion (Craw­ford et al. 1996b). [t appears, how· ever, that managed floods can be used ro partially restore riparian eco­systems.

It is essential to combine studies of ecosystem functional responses with community and population studies. By studying a broad spec­trum of ecological responses to man­aged floods, we hope to resolve em­pirically [he time course of ecosystem reorganization in response to hydro­logic manipulation. Moreover, or­ganizing experimental studies of res­toration around conceptual models is likely to lead to general principles for ecosystem restoration. For in­stance, our proposed model of eco­system reorganization led to tWO in­sights that should be considered during riparian restoration . Firsr, during restoration, the ecosystem cannot proceed directly from the dis­connected phase to the steady-state phase but must pass through a pe­riod of reorganization. Second, eco­system components reorganize at different rates, and re stora tion projects must recognize and consider this va riation in the tempo of reorga­nization. Out of this search for gen­eral ecological principles will come a better understanding of the struc­ture and function of riparian ecosys­tems and more effecti ve programs for their restoration.

Acknowledgments

Many people have contributed ro the studies described in this article. We thank Phil Norton, John Taylor,

September 1998

Mauged floods Ixgin ... 10,000 -r------'r------

Figure 6. The reorgani­zation model for ripar­ian restoration. Resto-

100

Rpmli"'" (mg C nil d·1 )

1 Sleaily·!UIe

ration involves three phases: the discon­nected phase, prior to the reestablishment of managed floods; the reo o rganization phase, in­duced by (he reestablish­ment of the flood pulse; and the steady-state phase, which follows the reo rganization phase. The predicted response

to restoration plotted here is forest -floor respiration (R). Dotted line includes proposed within-year variation; solid line presents predicted average annual rates.

and th e staff at Bosque del Apache a) National Wildlife Refuge for their gen­erous logistical sup­port. We are in­debted to Shivcharn Dhillion, Tom Kieft, and Carl White for help and advice on soils and soi l ecol-ogy and to numer-

Figure 7. Predicted and observed rates of forest ~fl oor respira­tion. (a) Conceptual model (from Figure 6). (b ) Observed rales of forest"f1oor respira tion at concrol sites, ar the exper im en tal flood site, and at the natural flood site.

b)

Managed floods begin .B. [0,000 ~----'=-------

100 R,.,.."",

(ma C ai1 d· l )

1 Discc:mtctfd Re- Sleady-sta~

orplizatioo

Time ,=*

.B. 10,000 .------"r---.,-----,

R Obstrvcd (1111 C ",," d·1 )

100

Dry Exp. Cootrols floods

NorunJ Floodo

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ous undergraduate and graduate stu­dents atthe University of New Mexico. Insightful discussions with Peter Jacobson were much appreciated. This research was partially supported by Cooperative Agreement 14-16-0002-91-228 between The Univer­sity of New Mexico and the US Fish and Wildlife Service, by NSF grant no. DEB- 9414767, and by NASA Ecosys­tern Restoration Award NAG5-6999.

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